United States

Environmental Protection

Agency

Air

EPA-452/R-97-004

December 1997

Mercury Study

Report to Congress

c7o032-1-1

Office of Air Quality Planning & Standards

and

Office of Research and Development

Volume II:

An Inventory of Anthropogenic

Mercury Emissions in the

United States
MERCURY STUDY REPORT TO CONGRESS

VOLUME II:

AN INVENTORY OF ANTHROPOGENIC MERCURY

EMISSIONS IN THE UNITED STATES

December 1997

Office of Air Quality Planning and Standards

and

Office of Research and Development

U.S. Environmental Protection Agency
i

TABLE OF CONTENTS

Page

U.S. EPA AUTHORS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . iii

SCIENTIFIC PEER REVIEWERS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . iv

WORK GROUP AND U.S. EPA/ORD REVIEWERS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . vii

LIST OF TABLES ................................................................. viii

LIST OF FIGURES .................................................................. x

LIST OF SYMBOLS, UNITS AND ACRONYMS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xi

EXECUTIVE SUMMARY ......................................................... ES-1

1. INTRODUCTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1-1

1.1 Overview of Sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1-1

1.2 Study Approach and Uncertainties . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1-2

1.3 Organization of the Rest of the Document . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1-4

2. TRENDS IN MERCURY CONSUMPTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-1

3. ANTHROPOGENIC AREA SOURCES OF MERCURY EMISSIONS . . . . . . . . . . . . . . . . . 3-1

3.1 Electric Lamp Breakage . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-1

3.2 General Laboratory Use . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-7

3.3 Dental Preparation and Use . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-7

3.4 Municipal Solid Waste Landfills . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-7

3.5 Mobile Sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-8

3.6 Paint Use . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-8

3.7 Agricultural Burning . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-9

3.8 Other Area Sources ................................................... 3-10

4. ANTHROPOGENIC POINT SOURCES OF MERCURY EMISSIONS . . . . . . . . . . . . . . . . . 4-1

4.1 Combustion Sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-1

4.1.1 Utility Boilers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-3

4.1.2 Municipal Waste Combustors ..................................... 4-15

4.1.3 Commercial/Industrial Boilers .................................... 4-26

4.1.4 Medical Waste Incinerators ...................................... 4-27

4.1.5 Hazardous Waste Combustors .................................... 4-30

4.1.6 Residential Boilers ............................................. 4-32

4.1.7 Sewage Sludge Incinerators ...................................... 4-33

4.1.8 Wood Combustion ............................................. 4-35

4.1.9 Crematories ................................................... 4-36

4.2 Manufacturing Sources ................................................ 4-36

4.2.1 Chlor-alkali Production Using the Mercury Cell Process ............... 4-36

4.2.2 Cement Manufacturing .......................................... 4-41

4.2.3 Pulp and Paper Manufacturing .................................... 4-43

4.2.4 Instrument (Thermometers) Manufacturing .......................... 4-45

4.2.5 Secondary Mercury Production ................................... 4-47

4.2.6 Electrical Apparatus Manufacturing ................................ 4-49

4.2.7 Carbon Black Production ........................................ 4-53
TABLE OF CONTENTS (continued)

Page

ii

4.2.8 Lime Manufacturing ............................................ 4-56

4.2.9 Primary Lead Smelting .......................................... 4-58

4.2.10 Primary Copper Smelting ........................................ 4-60

4.2.11 Fluorescent Lamp Recycling ..................................... 4-63

4.2.12 Battery Production ............................................. 4-64

4.2.13 Primary Mercury Production ..................................... 4-68

4.2.14 Mercury Compounds Production .................................. 4-70

4.2.15 Byproduct Coke Production ...................................... 4-71

4.2.16 Petroleum Refining ............................................. 4-73

4.3 Miscellaneous Sources ................................................. 4-74

4.3.1 Geothermal Power Plants ........................................ 4-74

4.3.2 Pigments, Oil Shale Retorting, Mercury Catalysts, Turf Products and

Explosives .................................................... 4-77

5. EMISSIONS SUMMARY . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-1

6. CONCLUSIONS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-1

7. RESEARCH NEEDS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-1

8. REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8-1

APPENDIX A INFORMATION ON LOCATIONS OF AND EMISSIONS FROM COMBUSTION

SOURCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A-1

APPENDIX B MERCURY REMOVAL CAPABILITIES OF PARTICULATE

MATTER AND ACID GAS CONTROLS FOR UTILITIES . . . . . . . . . . . . . . . . . B-1

APPENDIX C EMISSION MODIFICATION FACTORS FOR UTILITY BOILER

EMISSION ESTIMATES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C-1
iii

U.S. EPA AUTHORS

Principal Author:

Martha H. Keating

Office of Air Quality Planning and Standards

Research Triangle Park, NC

Contributing Authors:

Dennis Beauregard

Office of Air Quality Planning and Standards

Research Triangle Park, NC

William G. Benjey, Ph.D.

Atmospheric Sciences Modeling Division

Air Resources Laboratory

National Oceanic and Atmospheric Administration

Research Triangle Park, NC

on assignment to the

U.S. EPA National Exposure Research Laboratory

Laurel Driver

Office of Air Quality Planning and Standards

Research Triangle Park, NC

William H. Maxwell, P.E.

Office of Air Quality Planning and Standards

Research Triangle Park, NC

Warren D. Peters

Office of Air Quality Planning and Standards

Research Triangle Park, NC

Anne A. Pope

Office of Air Quality Planning and Standards

Research Triangle Park, NC
iv

SCIENTIFIC PEER REVIEWERS

Dr. William J. Adams* Elizabeth Campbell

Kennecott Utah Corporation U.S. Department of Energy

Dr. Brian J. Allee

Harza Northwest, Incorporated Dr. Rick Canady

Dr. Thomas D. Atkeson Registry

Florida Department of Environmental

Protection Dr. Rufus Chaney

Dr. Donald G. Barnes*

U.S. EPA Science Advisory Board Dr. Joan Daisey*

Dr. Steven M. Bartell

SENES Oak Ridge, Inc. Dr. John A. Dellinger*

Dr. David Bellinger*

Children’s Hospital, Boston Dr. Kim N. Dietrich*

Dr. Nicolas Bloom*

Frontier Geosciences, Inc. Dr. Tim Eder

Dr. Mike Bolger National Wildlife Federation for the

U.S. Food and Drug Administration States of Michigan and Ohio

Dr. Peter Botros Dr. Lawrence J. Fischer*

U.S. Department of Energy Michigan State University

Federal Energy Technology Center

Thomas D. Brown University of Connecticut

U.S. Department of Energy Avery Point

Federal Energy Technology Center

Dr. Dallas Burtraw* U.S. EPA Science Advisory Board

Resources for the Future

Dr. Thomas Burbacher* National Center for Health Statistics

University of Washington

Seattle Dr. Bruce A. Fowler*

Dr. James P. Butler

University of Chicago Dr. Steven G. Gilbert*

Argonne National Laboratory Biosupport, Inc.

Policy Office, Washington D.C.

Agency for Toxic Substances and Disease

U.S. Department of Agriculture

Lawrence Berkeley National Laboratory

Medical College of Wisconsin

University of Cincinnati

Great Lakes Natural Resource Center

Dr. William F. Fitzgerald

A. Robert Flaak*

Dr. Katherine Flegal

University of Maryland at Baltimore
SCIENTIFIC PEER REVIEWERS (continued)

v

Dr. Cynthia C. Gilmour* Dr. Genevieve M. Matanoski*

The Academy of Natural Sciences The Johns Hopkins University

Dr. Robert Goyer Dr. Thomas McKone*

National Institute of Environmental Health University of California

Sciences Berkeley

Dr. George Gray Dr. Malcolm Meaburn

Harvard School of Public Health National Oceanic and Atmospheric

Dr. Terry Haines U.S. Department of Commerce

National Biological Service

Dr. Gary Heinz* Wisconsin Department of Natural Resources

Patuxent Wildlife Research Center

Joann L. Held University of Texas Science Center at Houston

New Jersey Department of Environmental

Protection & Energy Dr. Paul Mushak

Dr. Robert E. Hueter* PB Associates

Mote Marine Laboratory

Dr. Harold E. B. Humphrey* U.S.Department of Energy

Michigan Department of Community Health Federal Energy Technology Center

Dr. James P. Hurley* Dr. Christopher Newland*

University of Wisconsin Auburn University

Madison

Dr. Joseph L. Jacobson* The University of Michigan

Wayne State University Ann Arbor

Dr. Gerald J. Keeler William O’Dowd

University of Michigan U.S. Department of Energy

Ann Arbor Federal Energy Technology Center

Dr. Ronald J. Kendall* Dr. W. Steven Otwell*

Clemson University University of Florida

Dr. Lynda P. Knobeloch*

Wisconsin Division of Health Dr. Jozef M. Pacyna

Dr. Leonard Levin

Electric Power Research Institute Dr. Ruth Patterson

Dr. Steven E. Lindberg* Fred Gutchinson Cancer Research Center

Oak Ridge National Laboratory

Administration

Dr. Michael W. Meyer*

Dr. Maria Morandi*

Harvey Ness

Dr. Jerome O. Nriagu*

Gainesville

Norwegian Institute for Air Research

Cancer Prevention Research Program
SCIENTIFIC PEER REVIEWERS (continued)

vi

Dr. Donald Porcella Dennis Smith

Electric Power Research Institute U.S. Department of Energy

Dr. Deborah C. Rice*

Toxicology Research Center Dr. Ann Spacie*

Samuel R. Rondberg*

U.S. EPA Science Advisory Board Dr. Alan H. Stern

Charles Schmidt Protection & Energy

U.S. Department of Energy

Dr. Pamela Shubat Earth Tech

Minnesota Department of Health

Dr. Ellen K. Silbergeld* Minnesota Pollution Control Agency

University of Maryland

Baltimore Dr. Valerie Thomas*

Dr. Howard A. Simonin*

NYSDEC Aquatic Toxicant Research Unit Dr. M. Anthony Verity

Federal Energy Technology Center

Purdue University

New Jersey Department of Environmental

Dr. David G. Strimaitis*

Dr. Edward B. Swain

Princeton University

University of California

Los Angeles

*With EPA’s Science Advisory Board, Mercury Review Subcommittee
vii

WORK GROUP AND U.S. EPA /ORD REVIEWERS

Core Work Group Reviewers: U.S. EPA/ORD Reviewers:

Dan Axelrad, U.S. EPA Robert Beliles, Ph.D., D.A.B.T.

Office of Policy, Planning and Evaluation National Center for Environmental Assessment

Angela Bandemehr, U.S. EPA

Region 5 Eletha Brady-Roberts

Jim Darr, U.S. EPA Cincinnati, OH

Office of Pollution Prevention and Toxic

Substances Annie M. Jarabek

Thomas Gentile, State of New York Research Triangle Park, NC

Department of Environmental Conservation

Arnie Kuzmack, U.S. EPA National Center for Environmental Assessment

Office of Water Washington, DC

David Layland, U.S. EPA Susan Braen Norton

Office of Solid Waste and Emergency Response National Center for Environmental Assessment

Karen Levy, U.S. EPA

Office of Policy Analysis and Review Terry Harvey, D.V.M.

Steve Levy, U.S. EPA Cincinnati, OH

Office of Solid Waste and Emergency Response

Lorraine Randecker, U.S. EPA

Office of Pollution Prevention and Toxic

Substances

Joy Taylor, State of Michigan

Department of Natural Resources

Washington, DC

National Center for Environmental Assessment

National Center for Environmental Assessment

Matthew Lorber

Washington, DC

National Center for Environmental Assessment
LIST OF TABLES

viii

Page

ES-1 Sources of Anthropogenic Mercury Emissions Examined in this Inventory ...... ES-3

ES-2 Anthropogenic Mercury Sources With Sufficient Data to Estimate

National Emissions .................................................. ES-5

ES-3 Best Point Estimates of 1994-1995 National Mercury Emission Rates by Category ES-6

1-1 Sources of Anthropogenic Mercury Emissions Examined in this Inventory . . . . . . . 1-3

1-2 Anthropogenic Mercury Sources With Sufficient Data to Estimate

National Emissions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1-5

1-3 Mercury Sources With Insufficient Information to Estimate

National Emissions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1-6

2-1 U.S. Mercury: Supply, Demand, Imports, Exports . . . . . . . . . . . . . . . . . . . . . . . . . . 2-2

3-1 Best Point Estimates of Mercury Emissions from Anthropogenic Area

Sources: 1994-1995 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-2

3-2 Mercury Content of Fluorescent Bulbs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-3

3-3 Mercury (HID) Lamp Production - 1970 to 1989 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-3

3-4 Mercury Content of HID Lamps . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-4

4-1 Best Point Estimates of Mercury Emissions from Combustion,

Manufacturing and Miscellaneous Point Sources: 1994-1995 . . . . . . . . . . . . . . . . . . 4-1

4-2 Best Point Estimates of Mercury Emissions from Anthropogenic Combustion

Point Sources: 1994-1995 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-2

4-3 Comparison of Mercury Concentrations in Raw and Cleaned Coal .............. 4-11

4-4 Best Point Estimate of Mercury Emissions from Utility Boilers: 1994-1995 ....... 4-12

4-5 Estimated Discards of Mercury in Products in Municipal Solid Waste ........... 4-19

4-6 Estimated Discards of Mercury in Batteries ................................ 4-22

4-7 Estimated Discards of Mercury in Paint Residues ........................... 4-24

4-8 Estimated Discards of Mercury in Thermostats ............................. 4-25

4-9 Best Point Estimate of Mercury Emissions from Anthropogenic Manufacturing 

Sources: 1994-1995 ................................................. 4-37

4-10 1996 U.S. Mercury-Cell Chlor-Alkali Production Facilities ................... 4-39

4-11 1995 Major U.S. Mercury Recyclers ..................................... 4-48

4-12 Discards of Mercury in Electric Switches .................................. 4-51

4-13 1995 U.S. Fluorescent Lamp Manufacturers' Headquarters ................... 4-52

4-14 1992 U.S. Carbon Black Production Facilities ............................. 4-54

4-15 Lime Producers in the U.S. in 1994 ...................................... 4-57

4-16 1994 U.S. Primary Lead Smelters and Refineries ........................... 4-58

4-17 1996 U.S. Primary Copper Smelters and Refineries .......................... 4-61

4-18 Mercury Ore Concentrate and Emissions from Primary Copper

Smelters in the U.S. .................................................. 4-62

4-19 1992 U.S. Mercuric Oxide, Alkaline Manganese, or ZincCarbon Button Cell Battery Manufacturers ................................ 4-65

4-20 Emission Source Parameters for an Integrated Mercury

Button Cell Manufacturing Facility ...................................... 4-67

4-21 1996 U.S. Byproduct Mercury-Producing Gold Mines ...................... 4-68

4-22 1995 U.S. Mercury Compound Producers. ................................ 4-70

4-23 1991 U.S. Byproduct Coke Producers .................................... 4-72
LIST OF TABLES (continued)

ix

Page

4-24 Best Point Estimates of Mercury Emissions from Miscellaneous

Anthropogenic Emission Sources: 1994-1995 ............................. 4-75

4-25 1992 U.S. Geothermal Power Plants ..................................... 4-76

4-26 Mercury Emission Factors for Geothermal Power Plants ..................... 4-77

5-1 Best Point Estimates of 1994-1995 National Mercury Emission Rates

by Category . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-2

5-2 Best Point Estimates of Mercury Emissions from Anthropogenic

Sources: 1994-1995 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-4

5-3 Mercury Area Sources Allocation Methodology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-5

A-1 1994 Mercury Emissions From Utility Boilers, By State and Fuel Type . . . . . . . . . . A-1

A-2 Estimates of 1994 Coal, Natural Gas, and Oil Consumption in the Commercial/

Industrial Sector Per State (Trillion Btu) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A-4

A-3 Estimates of Mercury Emissions From Coal-Fired Commercial/Industrial

Boilers on a Per-State Basis For 1994 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A-5

A-4 Estimates of Mercury Emissions From Oil-Fired Commercial/Industrial

Boilers On a Per-State Basis For 1994 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A-6

A-5 Estimates of 1994 Coal, Natural Gas, and Oil Consumption in the Residential

Sector Per State (Trillion Btu) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A-7

A-6 Estimates of Mercury Emissions From Coal-Fired Residential Boilers

on a Per-State Basis For 1994 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A-8

A-7 Estimates of Mercury Emissions From Oil-Fired Residential Boilers

on a Per-State Basis For 1994 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A-9

A-8 Existing MWC Facilities (As of 1995) ................................... A-10

A-9 Mercury Emissions From MWCs by Combustor Type For 1995 ............... A-15

A-10 MWI Population By State ............................................. A-16

B-1 Test Data for FGD Units . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B-3

B-2 Spray Dryer Adsorption Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B-5

B-3 Fabric Filter Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B-7

B-4 Test Data for Cold-Side Electrostatic Precipitators (Controlling

Coal-Fired Units) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B-9

B-5 Test Data for Hot-Side Electrostatic Precipitators (Controlling

Coal-Fired Units) .................................................... B-11

B-6 Test Data for Cold-Side Electrostatic Precipitators (Controlling

Oil-Fired Units) ..................................................... B-12

C-1 Emission Modification Factors for Utility Boiler Emission Estimates . . . . . . . . . . . . C-1
x

LIST OF FIGURES

Page

ES-1 Total 1994-95 U.S. Anthropogenic Mercury Emissions ...................... ES-8

2-1 U.S. Mercury: Supply, Demand, Secondary Production . . . . . . . . . . . . . . . . . . . . . . . 2-3

3-1 Overall Fate of Mercury from Used Mercury-Containing

Fluorescent Lamps . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-5

4-1 Location of Coal-Fired Utility Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-4

4-2 Location of Oil-Fired Utility Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-5

4-3 Comparison of Mercury Removal Efficiencies Without Activated Carbon

Injection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-7

4-4 Mercury Emissions from Oil- and Natural-Gas Fired Boilers .................. 4-13

4-5 Mercury Emissions from Coal-Fired Boilers ............................... 4-14

4-6 Municipal Waste Combustor Facilities .................................... 4-16

4-7 Discards of Mercury in Municipal Solid Waste, 1989 ........................ 4-20

4-8 Estimated Discards of Mercury in Electric Lighting in Municipal

Solid Waste ......................................................... 4-23

4-9 Estimated Discards of Mercury in Pigments in Municipal Solid Waste ........... 4-26

4-10 Sewage Sludge Incinerators ............................................. 4-34

4-11 Chlor-Alkali Production Facilities ........................................ 4-40

4-12 Cement Manufacturing Plants ........................................... 4-42

4-13 Carbon Black Manufacturing Facilities .................................... 4-55

4-14 Primary Lead Smelters ................................................ 4-59

4-15 Primary Copper Smelters ............................................... 4-61

4-16 1991 U.S. Byproduct Coke Producers ..................................... 4-71

5-1 Total 1994-95 U.S. Anthropogenic Mercury Emissions . . . . . . . . . . . . . . . . . . . . . . . . 5-6

B-1 Removal of Mercury By An FGD (Coal) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B-2

B-2 Removal of Mercury By A Spray Dryer Adsorber/Fabric Filter (Coal) . . . . . . . . . . . B-4

B-3 Removal of Mercury By A FF (Coal) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B-6

B-4 Removal of Mercury By Electrostatic Precipitators (Cold-Side, Coal) . . . . . . . . . . . B-8

B-5 Removal of Mercury By Electrostatic Precipitators (Hot-Side, Coal) ........... B-10

B-6 Removal of Mercury By Electrostatic Precipitators (Oil) ..................... B-11
xi

LIST OF SYMBOLS, UNITS AND ACRONYMS

AP-42 Compilation of Air Pollutant Emission Factors (U.S. EPA, 1988a)

APCD Air Pollution Control Device

Btu British Thermal Unit

CAA Clean Air Act as Amended in 1990

CAS Chemical Abstract Service

CFB Circulating fluidized bed

CFR Code of Federal Regulations

CIP Carbon-in-pulp process

COC Certification of Compliance

d Day

dscf Dry standard cubic foot

EEI Edison Electric Institute

EMF Emission modification factor

EPRI Electric Power Research Institute

ESP Electrostatic precipitator

FBC Fluidized bed combustor

FF Fabric filter

FGD Flue gas desulfurization

FTP Federal Test Procedure

g Gram

GW Gigawatt

HFET Highway Fuel Economy Test

Hg Mercury

HID High Intensity Discharge

hr Hour

ISGS Illinois State Geological Survey

kg Kilogram

kJ Kilojoules

L Liter

L&E Locating and Estimating Document (U.S. EPA, 1993a)

lb Pound

MB/REF Mass burn/refractory wall

MB/RC Mass burn/rotary waterwall

MB/WW Mass burn/water wall

Mg Megagram or metric ton (2200 pounds)

Mj Megajoules

mm Millimeter

MSW Municipal solid waste

MW Molecular weight

MWC Municipal waste combustor

MWI Medical Waste Incinerator

NEMA National Electrical Manufacturers Association

Nm Normal cubic meter 3

NSPS New Source Performance Standard

NYCC New York City Cycle

OPP U.S. EPA Office of Pesticides Programs
xii

OSHA Occupational Safety and Health Administration

OSW U.S. EPA Office of Solid Waste

PM Particulate matter

ppb Parts per billion

ppm Parts per million

ppmwt Parts per million by weight

RCRA Resource Conservation and Recovery Act

RDF Refuse derived fuel

SDA Spray dryer adsorber

SSI Sewage sludge incinerator

TRI Toxic Release Inventory

UDI Utility Data Institute

umol Micromole

USGS United States Geological Service

VOC Volatile Organic Compound

WDF Waste derived fuel

yr Year
ES-1

EXECUTIVE SUMMARY

Section 112(n)(1)(B) of the Clean Air Act (CAA), as amended in 1990, requires the U.S.

Environmental Protection Agency (U.S. EPA) to submit a study on atmospheric mercury emissions to

Congress. The sources of emissions that must be studied include electric utility steam generating units,

municipal waste combustion units and other sources, including area sources. Congress directed that the

Mercury Study evaluate many aspects of mercury emissions, including the rate and mass of emissions,

health and environmental effects, technologies to control such emissions and the costs of such controls.

In response to this mandate, U.S. EPA has prepared an eight-volume Mercury Study Report to

Congress. This volume -- Volume II of the Report to Congress -- estimates emissions of mercury from

anthropogenic sources and provides abbreviated process descriptions, control technique options,

emission factors and activity levels for these sources. The information contained in this volume will be

useful in identifying source categories that emit mercury, in selecting potential candidates for mercury

emission reductions and in evaluating possible control technologies or materials substitution/elimination

that could be used to achieve these reductions (as presented in Volume VIII of this Report to Congress). 

The emissions data presented here also served as input data to U.S. EPA's local impact analyses and

long-range transport model that assessed the dispersion of mercury emissions nationwide (as presented in

Volume III of this Report to Congress).

Overview of Sources

In the CAA, Congress directed U.S. EPA to examine sources of mercury emissions, including

electric utility steam generating units, municipal waste combustion units and other sources, including

area sources. The U.S. EPA interpreted the phrase "... and other sources..." to mean that a

comprehensive examination of mercury sources should be made and to the extent data were available, air

emissions should be quantified. This report describes in some detail various source categories that emit

mercury. In many cases, a particular source category is identified as having the potential to emit

mercury, but data are not available to assign a quantitative estimate of emissions. The U.S. EPA's intent

was to identify as many sources of mercury emissions to the air as possible and to quantify those

emissions where possible.

The mercury emissions data that are available vary considerably in quantity and quality between

different source types. Not surprisingly, the best available data are for source categories that U.S. EPA

has examined in the past or is currently studying.

Sources of mercury emissions in the United States are ubiquitous. To characterize these

emissions, the type of mercury emission is defined as either:

 Natural mercury emissions -- the mobilization or release of geologically bound mercury

by natural processes, with mass transfer of mercury to the atmosphere;

 Anthropogenic mercury emissions -- the mobilization or release of geologically bound

mercury by human activities, with mass transfer of mercury to the atmosphere; or

 Re-emitted mercury -- the mass transfer of mercury to the atmosphere by biologic and

geologic processes drawing on a pool of mercury that was deposited to the earth's

surface after initial mobilization by either anthropogenic or natural activities.
 The emission factors used in developing this mercury emissions inventory are generally consistent with those 1

presented in the U.S. EPA document entitled Locating and Estimating Air Emissions from Sources of Mercury and

Mercury Compounds (Draft Final Report) May 1997. Some of the nationwide emission estimates may vary slightly

between the two documents.

ES-2

Anthropogenic mercury emissions can be further divided into area and point sources. 

Anthropogenic area sources of mercury emissions are sources that are typically small and numerous and

usually cannot be readily located geographically. For the purpose of this report, mobile sources are

included in the area source discussion. Point sources are those anthropogenic sources that are associated

with a fixed geographic location. These point sources are further divided into combustion,

manufacturing and miscellaneous source categories. Particular types of sources that fall into these

various groups and that were examined in this study are outlined in Table ES-1.

A prerequisite for developing strategies for reducing mercury concentrations in surface waters,

biota and ambient air is a comprehensive characterization of all sources of mercury releases to the

environment. This would include a review not only of airborne emissions, but also direct discharges to

surface water and soil as well as past commercial and waste disposal practices (e.g., historical

applications of mercury-containing pesticides and fungicides that are presently banned) that have

resulted in mercury contamination of different environmental media. Although the focus of this study is

on air emissions in accordance with section 112(n) of the CAA, U.S. EPA recognizes that such past and

current releases of mercury to other media can be important contributors to overall mercury loadings and

exposures in some locations.

Moreover, a complete characterization of air emissions would include the identification of all

significant mercury emission sources, both anthropogenic and natural, and would account for re-emitted

mercury. The current state of knowledge about mercury emissions, however, does not allow for an

accurate assessment of either natural or re-emitted mercury emissions. For example, approximately onethird of total current global mercury emissions are thought to cycle from the oceans to the atmosphere

and back again to the oceans, but a major fraction of the emissions from oceans consists of recycled

anthropogenic mercury. It is believed that much less than 50 percent of the oceanic emission is from

mercury originally mobilized by natural sources. Similarly, an unknown but potentially large fraction of

terrestrial and vegetative emissions consists of recycled mercury from previously deposited

anthropogenic and natural emissions (Expert Panel, 1994).

Given the considerable uncertainties regarding the levels of natural and re-emitted mercury

emissions, this report focuses only on the nature and magnitude of mercury emissions from

anthropogenic sources. Further study is needed to determine the importance of natural and re-emitted

mercury.

Approach for Estimating Anthropogenic Emissions

For most anthropogenic source categories, an emission factor-based approach was used to

develop both facility-specific estimates for modeling purposes and nationwide emission estimates. This

approach requires an emission factor, which is a ratio of the mass of mercury emitted to a measure of

source activity. It also requires an estimate of the annual nationwide source activity level. Examples of 1

measures of source activity include total heat input for fossil fuel combustion and total raw material used

or product generated for industrial processes. Emission factors are generated from emission test data,

from engineering analyses based on mass balance techniques, or from transfer of information
ES-3

Table ES-1

Sources of Anthropogenic Mercury Emissions Examined in this Inventory

Area Point

Combustion Manufacturing Miscellaneous

Electric lamp breakage Utility Boilers Chlor-alkali production Oil shale retorting

Paints use Commercial/industrial Lime manufacturing Mercury catalysts

boilers

Laboratory use Residential boilers Primary mercury production Pigment production

Dental preparations Municipal waste Mercury compounds Explosives

combustors production manufacturing a

Mobile sources Medical waste incinerators Battery production Geothermal power a

plants

Agricultural burning Sewage sludge Electrical appartatus Turf products a

incinerators manufacturing

Landfills Hazardous waste Carbon black production

combustors

Sludge application Wood-fired boilers Byproduct coke production a a

Residential woodstoves Primary copper smelting a

Crematories Cement manufacturing

Primary lead smelting

Petroleum refininga

Instrument manufacturing

Secondary mercury

production

Zinc mininga

Fluorescent lamp recycling

Pulp and paper mills

Potential anthropogenic sources of mercury for which emissions were not estimated. a
ES-4

from comparable emission sources. Emission factors reflect the "typical control" achieved by the air

pollution control measures applied across the population of sources within a source category.

The emission factor-based approach does not generate exact emission estimates. Uncertainties

are introduced in the estimation of emission factors, control efficiencies and the activity level measures. 

Ideally, emission factors are based on a substantial quantity of data from sources that represent the

source category population. For trace pollutants like mercury, however, emission factors are frequently

based on limited data that may not have been collected from representative sources. Changes in

processes or emission measurement techniques over time may also result in biased emission factors. 

Emission control estimates are also generally based on limited data; as such, these estimates are

imprecise and may be biased. Further uncertainty in the emission estimates is added by the sources of

information used on source activity levels, which vary in reliability. Table ES-2 presents anthropogenic

source categories for which U.S. EPA had sufficient data to estimate national emissions.

Anthropogenic Emissions Summary

Table ES-3 summarizes the estimated national mercury emission rates by source category. 

While these emission estimates for anthropogenic sources have important limitations, they do provide

insight into the relative magnitude of emissions from different groups of sources. All of these emissions

estimates should be regarded as best estimates given available data.

Of the estimated 144 Megagrams (Mg) (158 tons) of mercury emitted annually into the

atmosphere by anthropogenic sources in the United States, approximately 87 percent is from combustion

point sources, 10 percent is from manufacturing point sources, 2 percent is from area sources, and

1 percent is from miscellaneous sources. Four specific source categories account for approximately

80 percent of the total anthropogenic emissions--coal-fired utility boilers (33 percent), municipal waste

combustion (19 percent), commercial/industrial boilers (18 percent), and medical waste incinerators

(10 percent). It should be noted that the U.S. EPA has finalized mercury emission limits for municipal

waste combustors and medical waste incinerators. When fully implemented, these emission limits will

reduce mercury emissions from these sources by an additional 90 percent over 1995 levels.

All four of the most significant sources represent high temperature waste combustion or fossil

fuel processes. For each of these operations, the mercury is present as a trace contaminant in the fuel or

feedstock. Because of its relatively low boiling point, mercury is volatilized during high temperature

operations and discharged to the atmosphere with the exhaust gas.

For the long-range transport analysis, the emissions inventory was mapped for the continental

U.S. The continental U.S. was divided into 40-km square grid cells and the magnitude of the mercury

emissions were calculated for each cell. For the most part, the location (at least to the county level) of

the mercury point sources described in this document were known.
ES-5

Table ES-2

Anthropogenic Mercury Sources With Sufficient

Data to Estimate National Emissions

Area Point

Combustion Manufacturing Miscellaneous

Electric lamp breakage Utility Boilers Chlor-alkali production Geothermal power

plants

Laboratory use Commercial/industrial Cement manufacturing

boilers

Dental preparation Residential boilers Battery production

Landfills Municipal waste Electric apparatus

combustors manufacturing

Medical waste Instrument manufacturing

incinerators

Sewage sludge Secondary mercury

incinerators production

Wood-fired boilers Carbon black production

Hazardous waste Primary lead smelting

combustors

Crematories Primary copper smelting

Lime manufacturing

Fluorescent lamp recycling

Pulp and paper mills
ES-6

Table ES-3

Best Point Estimates of 1994-1995 National Mercury Emission Rates by Category

Sources of mercury Mg/yr tons/yr Inventory a

1994-1995 1994-1995 % of Total bb b

Area sources 3.1 3.4 2.2

 Lamp breakage 1.4 1.5 1.0

 General laboratory use 1.0 1.1 0.7

 Dental preparations 0.6 0.7 0.4

 Landfills 0.07 0.08 0.1

 Mobile sources c c c

 Paint use c c c

 Agricultural burning c c c

Point Sources 141.0 154.7 97.8

 Combustion sources 125.3 137.7 86.9

 Utility boilers

 Coal

 Oil

 Natural gas

 MWCsh

 Commercial/industrial boilers

 Coal

 Oil

 MWIsh

 Hazardous waste combustors e

 Residential boilers

 Oil

 Coal

 SSIs

 Wood-fired boilersf

 Crematories

 Manufacturing sources

 Chlor-alkali

 Portland cement e

 Pulp and paper manufacturing

 Instruments manufacturing

 Secondary Hg production

 Electrical apparatus

 Carbon black

 Lime manufacturing

 Primary lead

 Primary copper

 Fluorescent lamp recycling

 Batteries

 Primary Hg production

 Mercury compounds 

 Byproduct coke

 Refineries

 Miscellaneous sources

 Geothermal power

 Turf products

 Pigments, oil, etc.

TOTAL

47.2 51.8 32.8

(47) (51.6) (32.6) d

(0.2) (0.2) (0.1)

(<0.1) (<0.1) (0.0)

26.9 29.6 18.7

25.8 28.4 17.9

(18.8) (20.7) (13.1)

(7.0) (7.7) (4.9)

14.6 16.0 10.1

6.4 7.1 4.4

3.3 3.6 2.3

(2.9) (3.2) (2.0)

(0.4) (0.5) (0.3)

0.9 1.0 0.6

0.2 0.2 0.1

<0.1 <0.1 0.0

14.4 15.6 10.0

6.5 7.1 4.5

4.4 4.8 3.1

1.7 1.9 1.2

0.5 0.5 0.3

0.4 0.4 0.3

0.3 0.3 0.2

0.3 0.3 0.2

0.1 0.1 0.1

0.1 0.1 0.1

<0.1 <0.1 0.0

<0.1 <0.1 0.0

<0.1 <0.1 0.0

ccc

ccc

ccc

ccc

1.3 1.4 0.9

1.3 1.4 0.9

ggg

ggg

144 158 100

 MWC=Municipal waste combustor; MWI=medical waste incinerator; SSI=sewage sludge incinerator a

Numbers do not add exactly because of rounding. b 

 Insufficient information to estimate 1994-1995 emissions. c

 Parentheses denote subtotal within larger point source category. d

For the purpose of this inventory, cement kilns that burn hazardous waste for fuel are counted as hazardous waste combustors. e 

 Includes boilers only; does not include residential wood combustion (wood stoves). f

 Mercury has been phased out of use. g

 EPA has finalized emissions guidelines for these source categories which will reduce mercury emissions by at least an additional 90 percent h

over 1995 levels. 
ES-7

Figure ES-1 illustrates the spatial distribution of mercury emissions across the U.S. based on this

inventory. This distribution formed the basis of the long-range transport modeling and the resulting

predictions of wet and dry deposition across the U.S.

Accuracy of the Inventory

The accuracy of the emission estimates is obviously a factor in assessing the inventory's

usefulness for its intended purposes. Considering the admitted gaps in the inventory, the external peer

review panel that reviewed this work in January 1995 concluded that the missing sources could

contribute as much as 20 percent more mercury emissions to the U.S. total. For comparison, one

reviewer submitted data on the amount of mercury emitted per person in some European countries (based

on anthropogenic emissions only).

Based on the inventory presented in this document, the U.S. inventory represents 0.55 g mercury

per person per year. Based on data submitted during the 1995 external peer review process, 0.90 g

mercury per person per year is emitted in the United Kingdom. In Germany (Western area), 0.75 g

mercury per person per year is emitted. In Poland, 0.88 g mercury per person per year is estimated to be

emitted. The European emission average is about 1.2 g mercury per person per year (Pacyna, 1995). 

This national inventory of estimated mercury emissions compares favorably with other national

estimates. Porcella, et al. (1995) estimated 1990 U.S. mercury emissions to be 154.1 Mg and Pai, et al.

(1997) estimated 1990 emissions at 146.4 Mg. This study estimates the 1994-1995 national baseline

emissions to be 145 Mg. In general, each of these studies used similar emissions estimation techniques

and data sources, and estimates for individual source categories are close. Like this study, these other

studies also used “top down” techniques based on emission factors (e.g., lbs mercury emitted per unit of

energy or lbs product produced) multiplied by an activity level (e.g., pounds product produced in a year). 

This approach is common, particularly for a national estimate where adding up actual emissions from

every source would be unrealistic.

A regional inventory being compiled by the Northeast States for Coordinated Air Use

Management (NESCAUM) was used for a regional modeling study of mercury emissions and dispersion

in Connecticut, Maine, Maryland, New Hampshire, New Jersey, New York, Rhode Island, and Vermont. 

Emissions for each state were allocated to modeling grid cells for regional modeling. A comparison of

the emissions inventory for each of these states to this study’s emission inventory for the same states

produced good agreement. The EPA’s emission inventory is about 19 Mg/year for the NESCAUM

states, while the states’ own estimates total about 16 Mg/year. The state estimates are likely to be more

accurate because in many cases, emissions testing is required for air pollution permits and these test data

were available to the states to estimate emissions from specific facilities (compared to the EPA’s

emission factor approach).

Trends in Mercury Emissions

It is difficult to predict with confidence the temporal trends in mercury emissions for the U.S.,

although there appears to be a trend toward decreasing total mercury emissions from 1990 to 1995. This

is particularly true for the waste combustion sources where emissions have declined 50 percent from

municipal waste combustors and 75 percent from medical waste incinerators since 1990 (see below). 

Also, as previously noted, there are a number of source categories where there is insufficient 

ES-9

data to estimate current emissions let alone potential future emissions. Based on available information,

however, a number of observations can be made regarding mercury emission trends from source

categories where some information is available about past activities and projected future activities.

Current emissions of mercury from manufacturing sources are generally low compared to

combustion sources (with the exception of chlor-alkali plants using the mercury cell process and portland

cement manufacturing plants). The emissions of mercury are more likely to occur when the product

(e.g., lamps, thermostats) is broken or discarded. Therefore, in terms of emission trends, one would

expect that if the future consumption of mercury remains consistent with the 1996 consumption rate,

emissions from most manufacturing sources would remain about the same.

For industrial or manufacturing sources that use mercury in products or processes, the overall

consumption of mercury is generally declining. Industrial consumption of mercury has declined by

about 75 percent between 1988 (1503 Mg) and 1996 (372 Mg). Much of this decline can be attributed to

the elimination of mercury as a paint additive (20 percent) and the reduction of mercury in batteries (36

percent). Use of mercury by other source categories remained about the same between 1988 and 1996.

Secondary production of mercury (i.e., recovering mercury from waste products) has increased

significantly over the past few years. While 372 Mg of mercury were used in industrial processes in

1996, 446 Mg were produced by secondary mercury producers and an additional 340 Mg were imported. 

This is a two-fold increase since 1991. The number of secondary mercury producers is expected to

increase as more facilities open to recover mercury from fluorescent lamps and other mercury-containing

products (e.g., thermostats). As a result there is potential for mercury emissions from this source

category to increase.

The largest identified source of mercury emissions during 1994-1995 is fossil fuel combustion

by utility boilers, particularly coal combustion. Future trends in mercury emissions from this source

category are largely dependent on both the nation's future energy needs and the fuel chosen to meet those

needs. Another factor is the nature of actions the utility industry may take in the future to meet other air

quality requirements under the Clean Air Act (e.g., national ambient air quality standards for ozone and

particulate matter).

Two other significant sources of mercury emissions currently are municipal waste combustors

and medical waste incinerators. Emissions from these source categories have declined considerably

since 1990 on account of plant closures (for medical waste incinerators) and reduction in the mercury

content of the waste stream (municipal waste combustors) and will decline even further by the year 2000

due to regulatory action the U.S. EPA is taking under the statutory authority of section 129 of the CAA. 

As described in sections 4.1.2 and 4.1.4 of this document, the U.S. EPA has finalized rules for municipal

waste combustors and medical waste incinerators that will, when fully implemented, reduce mercury

emissions from both of these source categories by an additional 90 percent over 1995 levels. In addition

to this federal action, a number of states (including Minnesota, Florida and New Jersey) have

implemented mandatory recycling programs to reduce mercury-containing waste, and some states have

regulations that impose emission limits that are lower than the federal regulation. These factors will

reduce national mercury emissions from these source categories even further.
ES-10

Conclusions

The following conclusions are presented in approximate order of degree of certainty in the

conclusion, based on the quality of the underlying database. The conclusions progress from

those with greater certainty to those with lesser certainty.

 Numerous industrial and manufacturing processes emit mercury to the atmosphere. 

Mercury emissions from U.S. manufacturing sources, however, have dropped about 75

percent over the past decade.

 Mercury is emitted, to a varying degree, from anthropogenic sources virtually

everywhere in the United States.

 Natural sources of mercury and re-emission of previously deposited mercury are also

sources of mercury to the atmosphere, although the magnitude of the contribution of

these sources relative to the contribution of current anthropogenic sources is not well

understood.

 Prior to 1995, municipal waste combustors and medical waste incinerators were the

largest identifiable source of mercury emissions to the atmosphere. Regulations which

have been finalized for municipal waste combustors and medical waste incinerators will,

when fully implemented, reduce emissions from these source categories by an additional

90 percent over 1995 levels.

 Present emissions estimates indicate that coal-fired utility boilers are the single largest

emissions source, contributing approximately 33 percent of the national inventory.

 Anthropogenic sources in the United States emit approximately 144 Mg (158 tons) of

mercury annually into the atmosphere. This estimate is believed to be accurate to within

30 percent. This estimate represents emissions calculated during the 1994-1995 time

frame. 

 In the United States, areas east of the Rocky Mountains have the highest concentration of

emissions from anthropogenic sources in the U.S.

 The areas having the greatest concentration of mercury emissions from anthropogenic

sources of total mercury (i.e., all chemical species) are the following: the urban corridor

from Washington D.C. to Boston, the Tampa and Miami areas of Florida, the larger

urban areas of the Midwest and Ohio Valley and two sites in northeastern Texas.

 The areas having generally the lowest emissions are in the Great Basin region of the

western United States and the High Plains region of the central United States. There are

generally few large emission sources in the western third of the United States, with the

exception of the San Francisco and Los Angeles areas and specific industrial operations.
ES-11

There are many uncertainties in the emission estimates for individual source categories due to

uncertainties inherent in an emission factor approach. The source of these uncertainties include

the following: 

 Variability in the estimates of source activity for each source category. Activity levels

used in this Report were compiled over different time periods and by a variety of survey

procedures.

 Emissions test data that are of poor quality or are based on very few analyses, which may

not be representative of the full source population being studied.

 Changes in processes or emission measurement techniques over time (especially since

about 1985). Earlier techniques may have measured too much mercury because of

contamination problems.

 A lack of data for some source categories which either led to estimates based on

engineering judgment or mass balance calculations. For a number of source categories

there were insufficient data and, thus, no emissions estimates were made.

 Limited data on the effectiveness of air pollution control equipment to capture mercury

emissions.

Understanding the public health and environmental impacts of current anthropogenic emissions

is complicated by an incomplete understanding of the following factors:

 Global and transboundary deposition of mercury and the impact this has on deposition of

mercury in the U.S.

 The magnitude and chemical nature of natural emissions.

 The magnitude and chemical nature of re-emitted mercury.

 The public health and environmental impacts of emissions from past uses of mercury

(such as paint application) relative to current anthropogenic emissions. 

To improve the emissions estimates, U.S. EPA would need the following:

 Source test data from a number of source categories that have been identified in this

volume as having insufficient data to estimate emissions. Notable among these are

mobile sources, agricultural burning, sludge application, coke ovens, petroleum refining,

residential woodstoves, mercury compounds production and zinc mining.

 Improvements in the existing emissions information for a number of source categories

including secondary mercury production (i.e., recycling), commercial and industrial

boilers, landfills, electric lamp breakage, and iron and steel manufacturing.

 Validation of a stack test protocol for speciated mercury emissions.
ES-12

 More data on the efficacy of conventional coal cleaning and the potential for slurries

from the cleaning process to be a mercury emission source.

 More data are needed on the mercury content of various coals and petroleum and the

trends in the mercury content of coal burned at utilities and petroleum refined in the U.S.

 Additional research to address the potential for methylmercury to be emitted (or formed)

in the flue gas of combustion sources.

 Investigation of the importance (quantitatively) of re-emission of mercury from

previously deposited anthropogenic emissions and mercury-bearing mining waste. This

would include both terrestrial and water environments. Measuring the flux of mercury

from various environments would allow a determination to be made of the relative

importance of re-emitted mercury to the overall emissions of current anthropogenic

sources.

 Determination of the mercury flux from natural sources to help determine the impact of

U.S. anthropogenic sources on the global mercury cycle as well as the impact of all

mercury emissions in the United States.

 More detailed emissions data to support the use of more sophisticated fate and transport

models for mercury; in particular, more information is needed on the chemical species of

mercury being emitted (including whether these species are particle-bound) and the

temporal variability of the emissions.

Based on trends in mercury use and emissions, the U.S. EPA predicts the following:

 A significant decrease (at least 90 percent over 1995 levels) will occur in mercury

emissions from municipal waste combustors and medical waste incinerators by the year

2000 when the regulations finalized by U.S. EPA for these source categories are fully

implemented.

 Manufacturing use of mercury will continue to decline with chlorine production from

mercury cell chlor-alkali plants continuing to account for most of the mercury use in the

manufacturing sector.

 Secondary production of mercury will continue to increase as more recycling facilities

commence operation to recover mercury from discarded products and wastes.
1-1

1. INTRODUCTION

Section 112(n)(1)(B) of the Clean Air Act (CAA), as amended in 1990, requires the U.S.

Environmental Protection Agency (U.S. EPA) to submit a study on atmospheric mercury emissions to

Congress. The sources of emissions that must be studied include electric utility steam generating units,

municipal waste combustion units and other sources, including area sources. Congress directed that the

Mercury Study evaluate many aspects of mercury emissions, including the rate and mass of emissions,

health and environmental effects, technologies to control such emissions, and the costs of such controls.

In response to this mandate, U.S. EPA has prepared an eight-volume Mercury Study Report to

Congress. The eight volumes are as follows:

I. Executive Summary

II. An Inventory of Anthropogenic Mercury Emissions in the United States

III. Fate and Transport of Mercury in the Environment

IV. An Assessment of Exposure to Mercury in the United States

V. Health Effects of Mercury and Mercury Compounds

VI. An Ecological Assessment for Anthropogenic Mercury Emissions in the United States

VII. Characterization of Human Health and Wildlife Risks from Mercury Exposure in the

United States

VIII. An Evaluation of Mercury Control Technologies and Costs

This volume (Volume II) estimates mercury emissions from anthropogenic sources and provides

abbreviated process descriptions, control technique options, emission factors, and activity levels for these

sources. Also, if sufficient information is available, locations by city, county, and state are given for

point sources.

1.1 Overview of Sources

In the CAA, Congress directed U.S. EPA to examine sources of mercury emissions, including

electric utility steam generating units, municipal waste combustion units and other sources, including

area sources. The U.S. EPA interpreted the phrase "... and other sources..." to mean that a

comprehensive examination of mercury sources should be made and to the extent data were available, air

emissions should be quantified. This report describes in some detail various source categories that emit

mercury. In many cases, a particular source category is identified as having the potential to emit

mercury, but data are not available to assign a quantitative estimate of emissions. The U.S. EPA's intent

was to identify as many sources of mercury emissions to the air as possible and to quantify those

emissions where possible.

The mercury emissions data that are available vary considerably in quantity and quality between

different source types. Not surprisingly, the best available data are for source categories that U.S. EPA

has examined in the past or is currently studying.

Sources of mercury emissions in the United States are ubiquitous. To characterize these

emissions, the type of mercury emission is defined as either:

 Natural mercury emissions -- the mobilization or release of geologically bound mercury

by natural processes, with mass transfer of mercury to the atmosphere;
1-2

 Anthropogenic mercury emissions -- the mobilization or release of geologically bound

mercury by human activities, with mass transfer of mercury to the atmosphere; or

 Re-emitted mercury -- the mass transfer of mercury to the atmosphere by biologic and

geologic processes drawing on a pool of mercury that was deposited to the earth's

surface after initial mobilization by either anthropogenic or natural activities.

Anthropogenic mercury emissions can be further divided into area and point sources. Anthropogenic

area sources of mercury emissions are sources that are typically small and numerous and usually cannot

be readily located geographically. For the purpose of this report, mobile sources are included in the area

source section. Point sources are those anthropogenic sources that are associated with a fixed geographic

location. These point sources are further divided into combustion, manufacturing and miscellaneous

source categories. Particular types of sources that fall into these various groups are outlined in Table 1-

1.

A prerequisite for developing strategies for reducing mercury concentrations in surface waters,

biota and ambient air is a comprehensive characterization of all sources of mercury releases to the

environment. A complete characterization would include: (1) all sources of airborne emissions,

including natural and anthropogenic emissions as well as re-emitted mercury; (2) direct discharges to

surface water and soil; and (3) past commercial and waste disposal practices that have resulted in

mercury contamination in different environmental media. The focus of this study, however, is only on

air emissions in accordance with section 112(n) of the CAA. In addition, the current state of knowledge

about airborne emissions does not allow for an accurate assessment of either natural mercury emissions

or re-emitted mercury. The U.S. EPA recognizes that an assessment of the relative public health and

environmental impact that can be attributed to current anthropogenic emissions is greatly complicated by

releases to other media, natural mercury emissions, and previous emissions of mercury that have

subsequently deposited. This report provides the basis for a nationwide mercury emission

characterization from anthropogenic sources. For each source category, the processes yielding mercury

emissions and the emission control measures are described. The procedures used to estimate nationwide

mercury emissions from each category are also delineated.

1.2 Study Approach and Uncertainties

This report contains mercury emission factors available from the U.S. EPA document Locating

and Estimating Air Emissions from Sources of Mercury and Mercury Compounds (L&E document, U.S.

EPA, 1997a). Other information sources used include recently published reports, journal articles and

information from trade associations. Mercury emission rates presented in this report are estimates only. 

To the degree that information is available, the sources of uncertainty in the emission estimates are

discussed (at least qualitatively) as the estimates are discussed throughout the report.

For most source categories, an emission factor-based approach was used to calculate nationwide

emission rate estimates. This approach requires an emission factor, which is a ratio of the mass of

mercury emitted to a measure of source activity, as well as an estimate of the annual nationwide source

activity level. Examples of measures of source activity include total heat input for fossil fuel combustion

and total raw material used or product generated for industrial processes. Activity levels used in this

report were compiled over different time periods and with a variety of survey procedures. Emission

factors are generated from emission test data, engineering analyses based 
1-3

Table 1-1

Sources of Anthropogenic Mercury Emissions Examined In This Inventory

Area Point

Combustion Manufacturing Miscellaneous

Electric lamp breakage Utility boilers Chlor-alkali production Oil shale retorting

Paint use Commercial/industrial Lime manufacturing Mercury catalysts

Laboratory use Primary mercury production Geothermal power plants

Dental preparations Mercury compounds Municipal waste landfills

Mobile sources combustors a

Agricultural burning Medical waste incinerators a

Landfills Sewage sludge incinerators manufacturing

Sludge application Hazardous waste Carbon black production a

boilers

Residential boilers

Municipal waste production

combustors

Wood-fired boilers

Residential

woodstoves Cement manufacturing a

Crematories Primary lead smelting

a

Battery production

Electrical apparatus

Byproduct coke production

Primary copper smelting

Petroleum refininga

Instrument

manufacturing

Secondary mercury

production

Zinc mininga

Fluorescent lamp recycling

Pulp and paper mills

 Potential anthropogenic sources of mercury for which emissions were not estimated. a
1-4

on mass balance techniques, or transfer of information from comparable sources. Generally, emission

factors are based on a limited set of test data that may not be representative of the full source population

being studied. Emission factors used to estimate nationwide emissions reflect "typical control" achieved

by the air pollution control measures applied across the population of sources within a source category. 

The emission factors and control levels used to develop emission estimates contained in this report were

generally taken from the L&E document (U.S. EPA, 1997a). Emission factors from the L&E document

were not used for estimating emissions from utility boilers, chlor-alkali plants using the mercury cell

process or fluorescent lamp breakage. Additional test data for utility boilers became available after the

L&E document was published. More recent information was also available directly from chlor-alkali

plant managers. A mass balance approach was used for lamp breakage.

The emission factor-based approach does not generate exact nationwide emission estimates. 

Uncertainties are introduced in the emission factors, the estimates of control efficiency and the

nationwide activity level measures. Ideally, emission factors are based on a substantial quantity of data

from sources that represent the source category population. For trace pollutants like mercury, however,

emission factors are frequently based on limited data that may not have been collected from

representative sources. Also, changes in processes or emission measurement techniques over time may

result in biased emission factors. In particular, analytical methods for detecting mercury have changed,

especially since about 1985. Emission control estimates are also generally based on limited data; as such

these estimates are imprecise and may be biased. Control efficiencies based on data collected using

older test methods may be biased because the older test methods tended to collect mercury vapor

inefficiently. (Currently, U.S. EPA Method 301 from 40 CFR Part 63, Appendix A can be used to

validate the equivalency of new methods.) Finally, activity levels used in this study were based on the

most recent information that was readily available. The sources of data used vary in reliability, adding

further uncertainty to the emission estimates.

Generally, quantitative estimates of the uncertainty in the emission factors, control efficiency

estimates and activity level measures are not available; these uncertainties are discussed qualitatively. 

Potential biases in the final emission estimates are also discussed. Table 1-2 presents source categories

for which U.S. EPA had sufficient data to estimate national emissions. Table 1-3 presents source

categories for which information is insufficient to estimate national emissions.

1.3 Organization of the Rest of the Document

The remainder of this volume consists of seven chapters and three appendices. Chapter 2

discusses trends in the environmental mercury burden and in the industrial consumption of mercury. 

Chapter 3 characterizes mercury emissions from area sources such as engines, light bulbs and dental

preparations. It describes the emitting process and presents the basis for the emission estimates. Chapter

4 provides a summary of emission estimates from point sources, including combustion, manufacturing

and miscellaneous sources. Chapter 5 summarizes mercury emission estimates from area and point

sources; Chapter 6 presents overall conclusions; Chapter 7 identifies further research needs; and all of

the references used are listed in Chapter 8. Appendix A contains information on activity levels, source

locations and emissions from some source categories. Appendix B presents available data on the

mercury removal efficiencies of particulate matter and acid gas controls for utilities. Finally, Appendix

C presents emission factors used to estimate emissions from utility boilers.
1-5

Table 1-2

Anthropogenic Mercury Sources With Sufficient

Data to Estimate National Emissions

Area Point

Combustion Manufacturing Miscellaneous

Electric lamp breakage Utility Boilers Chlor-alkali production Geothermal power

plants

Laboratory use Commercial/industrial Cement manufacturing

boilers

Dental preparation Residential boilers Battery production

Landfills Municipal waste Electric apparatus

combustors manufacturing

Medical waste Instrument manufacturing

incinerators

Sewage sludge Secondary mercury

incinerators production

Wood-fired boilers Carbon black production

Hazardous waste Primary lead smelting

combustors

Crematories Primary copper smelting

Lime manufacturing

Fluorescent lamp recycling

Pulp and paper mills
1-6

Table 1-3

Mercury Sources With Insufficient Information to Estimate National Emissions

Natural Anthropogenic

Area Point

Combustion Manufacturing Miscellaneous

Oceans and Mobile sources Residential Primary mercury Oil shale retorting

other natural woodstoves production

waters Paint use Mercury catalysts

Vegetation Agricultural production Pigment production

Volcanoes Petroleum refining Explosives

Rocks application Zinc mining

Soils

Wildfires

a

burning

Sludge manufacturing

a

Mercury compounds

a

a

a

a

Turf productsa

 Mercury is no longer used in U.S. manufacture. However, this is not meant to imply that these previous activities are no longer a

having an impact on the environment due to mercury's persistence in the environment.
2-1

2. TRENDS IN MERCURY CONSUMPTION

The mercury available for use in the U.S. comes from five main sources: (1) primary production

(mining); (2) by-product production (i.e., mercury by-product from gold mining operations); (3)

secondary production (recovery) from industrial recycling operations; (4) sales from excess government

stocks, including those held by the Department of Energy (DOE) and the Defense Logistics Agency

(DLA) within the Department of Defense; and (5) imports. Table 2-1 illustrates the relative

contributions of these sources to the U.S. mercury supply from 1988 through 1996. The table also shows

the total industrial demand or consumption levels for that same period.

Figure 2-1 plots mercury supply and demand levels since 1955. Supplies associated with byproduct production are not shown in this figure because data for this category are not available prior to

1990. Similarly, DLA sales are not presented in Figure 2-1 because data for such sales are not available

prior to 1982.

These data show a general decline in domestic mercury use since demand peaked in 1964. 

Domestic demand fell by 74 percent between 1980 and 1993, and by more than 75 percent between 1988

and 1996. The rate of decline, however, has slowed since 1990. Further evidence of the declining need

for mercury in the U.S. is provided by the general decline in imports since 1988 and the fact that exports

have exceeded imports since at least 1989. Federal mercury sales steadily increased from 1988 to 1993,

reaching a peak of 97 percent of the domestic demand. However, in July 1994, DLA suspended future

sales of mercury from the Department of Defense stockpile until the environmental implications of these

sales are addressed. In addition, in past years, DLA sold mercury accumulated and held by the

Department of Energy, which is also considered excess to government needs. DLA suspended these

mercury sales in July 1993 for an indefinite period in order to concentrate on selling material from its

own mercury stockpile (Ross & Associates, 1994). These suspensions caused federal sales to rapidly

decrease to 18 percent in 1994 and to zero since 1995 (Plachy, 1997).

In general, these data suggest that industrial manufacturers that use mercury are shifting away

from mercury except for uses for which mercury is considered essential. This shift is believed to be

largely the result of federal bans on mercury additives in paint and pesticides; industry efforts to reduce

mercury in batteries; increasing state regulation of mercury emissions sources and mercury in products;

and state-mandated recycling programs. A number of federal activities are also underway to investigate

pollution prevention measures and control techniques for a number of sources categories (see Volume

VIII of this Report to Congress).
2-2

Table 2-1

U.S. Mercury: Supply, Demand, Imports, Exports

(Mg)

Category 1988 1989 1990 1991 1992 1993 1994 1995 1996

Supply:

Mine production 379 414 448 0 0 0 0 0 0 a

By-product production W W 114 58 64 W W W W b c

Industrial recovery 278 137 108 165 176 350 466 534 446

DLA sales 52 170 52 103 267 543 86 0 0

DOE sales 214 180 193 215 103 0 0 0 0

Subtotal: federal sales 266 350 245 318 370 543 86 0 0

Imports 329 131 15 56 92 40 129 377 340

Total supply 1,252 1,032 930 597 702 933 681 911 786

Federal sales as % of 21.2% 33.9% 26.3% 53.3% 52.7% 58.2% 12.6% 0.0% 0.0%

total supply

Demand: 1,503 1,212 720 554 621 558 483 436 372

Federal sales as % of 17.6% 28.9% 34% 57.4% 59.6% 97.3% 17.8% 0.0% 0.0%

domestic demand

Imports: 329 131 15 56 92 40 129 377 340

Exports: N/A 221 311 786 977 389 316 179 45 d

Exports minus

imports:

N/A 90 296 730 885 349 187 -198 -295

Source: Plachy, 1997.

 Mercury production from McDermitt mine; closed November 1990. a

 Mercury by-product from nine gold mining firms. b

 Withheld for proprietary reasons. c

 Not available. d

3-1

3. ANTHROPOGENIC AREA SOURCES OF MERCURY EMISSIONS

Area sources account for approximately 2.2 percent of mercury emissions from anthropogenic

sources. Table 3-1 summarizes the estimated annual quantities of mercury emitted from area sources. 

Each of these source categories is discussed in turn in the sections that follow.

3.1 Electric Lamp Breakage

Electric lamps containing mercury include fluorescent, mercury vapor, metal halide and highpressure sodium lamps. More than half a billion mercury-containing lamps are produced each year

(O’Connell, 1997). These lamps are used for both indoor and outdoor applications including heat lamps,

lights for high-ceiling rooms, film projection, photography, dental exams, photochemistry, water

purification and street lighting. When these electric lamps are broken during use or disposal, a portion of

the mercury contained in them is emitted to the atmosphere. The amount of mercury emitted to the

atmosphere when mercury-containing lamps are disposed of will be a function of many factors. These

include the chemical form of mercury in the lamp and the size of the particulate forms of mercury in the

lamp powder. Approximately 643 Mg of mercury were discarded in U.S. municipal solid waste (MSW)

in 1989. The amount of mercury entering the MSW system from the disposal of used mercurycontaining lamps in 1989 is estimated to have been 24.3 Mg (26.8 tons), or 3.8 percent of the total

mercury content of MSW (Truesdale et al., 1993).

Mercury emissions due to lamp breakage are expected to decrease in the future for a number of

reasons. One reason is that states are beginning to view recycling as a viable option to decrease mercury

emissions. There is presently a bill in Massachusetts that would require every manufacturer of mercurycontaining products that may be sold or offered for sale to ensure that proper recycling of these products

occurs by funding a collection system. In addition, there have been technological advancements in the

manufacture of fluorescent lamps. Philips Lighting has devised a method to produce fluorescent lamps

with low-mercury technology which contain less than 10 mg of mercury per lamp. The company has

pledged that 80 percent of all its lamps sold in the United States will feature this technology by the end

of 1997 (O’Connell, 1997). The combination of increased regulation and advanced technology are

expected to have a significant impact on the amount of mercury that enters the MSW due to lamp

breakage.

Since the mid-1980s, electrical manufacturers have reduced the average amount of mercury in

each fluorescent bulb from an average of 48.2 mg to an average of 22.8 mg of mercury. A certain

amount of mercury is needed, however, in order to maintain desirable properties. The present practical

limit needed for full-rated life performance of a 4-foot fluorescent lamp has been thought to be about 15

mg of mercury (National Electrical Manufacturers Association, 1995). However, as noted above, Philips

Lighting recently announced that it will be manufacturing four-foot lamps with less than 10 mg of

mercury by late 1995 (Walitsky, 1995). Table 3-2 presents the estimated mercury content of fluorescent

bulbs, as provided in four different sources.

The average lifetime of a High Intensity Discharge (HID) lamp is between 10,000 and 24,000

hours. (Some small volume specialty products have lifetimes less than 10,000 hours or greater than

24,000 hours.) HID lamps last three to six years in typical applications. Low-pressure fluorescent lights

typically have a rated lifetime of 20,000 hours (Truesdale et al., 1993).

Approximately 550 million lamps containing mercury are sold annually in the United States

(National Electrical Manufacturers Association, 1992). Of these, 22 million are of the HID variety; the
3-2

Table 3-1

Best Point Estimates of Mercury Emissions from Anthropogenic Area Sources: 1994-1995

Source Data Uncertainty Basis for Emis

Mg/yr Tons/yr % of total

Emissions

Date of Degree of

a b

Electric lamp breakage 1.4 1.5 1.0 1989/1989 High Industry estimate for th

0.18 tons/year; this diffSection 3.1

Laboratory use 1.0 1.1 0.7 1973/1994 High Engineering judgmentDental preparations 0.6 0.7 0.4 1981/1995 High Engineering judgmentLandfills 0.07 0.08 0.1 1996/1995 High Test dataMobile sources - - - - - Insufficient information 

emissions

Paints use - - - - - Mercury phased out of Agricultural burning - - - - - Insufficient information 

emissions; one study es(0.04 tons/yr) from prehsugarcane in Florida ev

Total 3.1 3.4 2.2

 Date that data emission factor is based on/Date of activity factor used to estimate emissions. a

 A "medium" degree of uncertainty means the emission estimate is believed to be accurate within + 25 percent. A "high" degree of uncertainty means the emission estimate is believed to 

b
3-3

Table 3-2

Mercury Content of Fluorescent Bulbsa

Average Mercury Content (mg) per Bulb

Year NEMA CWF 3M

1970-1984 75

1985-1989 48.2 55

1990 41.6

1992 40 15-30

1995 22.8b

 Cole et al., 1992; National Electrical Manufacturers Association, 1992; Tanner, 1992; National Electrical Manufacturers a

Association, 1995.

 Philips Lighting has devised a method to produce fluorescent lamps with low-mercury technology which contain less b

than 10 mg of mercury per lamp.

Table 3-3

Mercury (HID) Lamp Production - 1970 to 1989a

Year Quantity (1000 bulbs) Year Quantity (1000 bulbs) b b

1970 6,841 1982 20,891

1971 7,684 1983 22,146

1972 8,420 1984 25,636

1973 9,349 1985 25,529

1974 9,158 1986 22,206

1975 8,737 1987 28,143

1976 10,383 1988 24,479

1977 10,853 1989 28,090

1978 12,175

1979 13,532

1980 30,187

1981 21,397

 Cole et al., 1992; U.S. EPA, 1992a. a

 Production rate = Domestic shipments - Exports + Imports. b
3-4

remaining 528 million are fluorescent bulbs. Table 3-3 contains production rates from 1970 through

1989 including exports and imports. Since 1970, there has been an increase in the production of HID

lamps (U.S. EPA, 1992a). Table 3-4 presents the mercury content of HID lamps and their

manufacturers.

Mercury and metal halide lamps consist of an inner quartz arc tube enclosed in an outer envelope

of heat resistant glass. The quartz arc tube contains a small amount of mercury ranging from 20 mg in a

75-watt lamp up to 250 mg in a 1000 watt lamp. According to the National Electrical Manufacturers

Association, no other substance has been found to replace mercury. High-pressure sodium lamps consist

of an inner, high-purity alumina ceramic tube enclosed in an outer envelope of heat-resistant glass. The

ceramic tube contains a small amount of sodium/mercury amalgam, ranging from 8.3 mg of mercury in a

50-watt lamp up to 25 mg in a 1000-watt lamp (National Electrical Manufacturers Association, 1992).

Table 3-4

Mercury Content of HID Lampsa

Manufacturer Type Mercury Content (mg)

Philips 250 watt HID 45

400 watt HID 60

1000 watt HID 70

Sylvania 250 watt HID 46

400 watt HID 75

1000 watt HID 75

 Cole et al., 1992; U.S. EPA, 1992a. a

The fate of used lamps is tied to the disposal of MSW. The three primary options for MSW

disposal are land filling, combustion and recycling. Land filling accounts for 82 percent of MSW

disposal, incineration accounts for 16 percent and recycling accounts for 2 percent. One study traced the

path of used lamps in MSW to each of the primary disposal options. Figure 3-1 diagrams the flow of

used mercury-containing lamps through the national MSW management system. 

On July 27, 1994, the US EPA published a proposed rule addressing the management of spent

mercury-containing lamps (59 FR 39288). In the proposal, the Agency presented two options for

changing the regulations governing spent mercury-containing lamps: 1) to add mercury-containing lamps

to the universal waste regulations, which would require special handling procedures to minimize lamp

breakage and disposal at designated sites (subject to RCRA hazardous waste regulations), or 2) to

conditionally exempt mercury-containing lamps from regulation as hazardous waste and require disposal

at EPA-permitted municipal solid waste landfills or a registered mercury reclamation facility, and record

keeping by generators.
To Atmosphere

4.0 (16.5%)

Fly Ash

0.14 (0.6%)

Bottom Ash

0.14 (0.6%)

Landfill

19.8 (81%)

Flue Gas

2.6(11%)

Incineration

2.85 (12%)

To Atmosphere

1.4 (5.9%)

To Atmosphere

0.01 (0.04%)

Air Emissions

0.005 (0.02%)

Recycle

0.48 (2%)

Recovery

or Disposal

0.47(2%)

Subtitle D

Waste

Management*

22.4 (92%)

Transport

23.8 (98%)

Transport

& Storage

0.49 (2%)

Storage,

Transport

and

Disposal

Mercury

in

Waste Lamps

24.3 Mg

98%

98%

92%

2%

to recycling

2%

6% from breakage

13%

87%

1% 99%

AirEmissions

(insufficient data toquantify)

90%

5%

5%

( ) = % of total (24.3Mg)

Contributions of mercury to atmosphere

from lamp breakage.

Figure 3-1

Overall Fate of Mercury from

Used Mercury-Containing Fluorescent Lamps

* It should be noted that some lamps in the municipal waste stream may go to subtitle C (hazardous waste) management.

This portion is not followed here and would be included in this Subtitle D waste stream.

Reference: Adapted from Truesdale et al., 1993.

3-5
3-6

EPA’s Office of Solid Waste modeled anticipated mercury emissions under these options, taking

into account any potential differences in lamps purchased by commercial establishments or changes in

utility power usage (including mercury emitted from utility power plants). EPA found that under either

option, the contribution of mercury emissions from landfills would be minimal. This is largely because,

based on model data, most lamps are broken before being land filled. Secondly, the Agency believes the

mix of lamp types purchased by commercial establishments would be independent of the option chosen. 

Taken collectively, these observations suggest that, to reduce lamp mercury emissions under either

option, procedures should be established that minimize emissions during transport and/or processing

(e.g., crushing) of spent lamps (U.S. EPA, 1997b).

 Ninety-eight percent of used lamps are managed as MSW under Subtitle D (the solid, nonhazardous waste program) of the Resource Conservation and Recovery Act (RCRA), with the remaining

2 percent being recycled. Mercury emissions from lamp breakage occur during transportation and

storage of lamps. A total of 1.4 Mg/yr (1.5 tons/yr) is estimated to be emitted during transport and

storage (Truesdale et al., 1993), as explained below. Additional mercury 

emissions from electric lamps are associated with MSW incineration, lamp recycling activities and

landfills. Mercury emissions from MSW incineration are accounted for in Section 4.1. Lamp recycling

activities are discussed in Section 4.2.7. An estimate of mercury emissions from landfills is found in

Section 3.7.

Discarded lamps may be transported in two ways: in garbage trucks as household or commercial

trash and in closed vans or trailers as part of a bulk re-lamping program. Of the 98 percent of mercury

from lamps in the MSW stream, 80 percent is transported in garbage trucks along with other solid waste

and 20 percent is transported in group re-lamping trucks holding lamps alone. Emissions from both

transport mechanisms were estimated using the waste pile mass transfer model developed for the RCRA

air emissions standards.

For transportation in a garbage truck, it was assumed that all lamps are broken in the truck and

that all of the mercury vapor is emitted to the atmosphere. The mercury concentration in the lamps was

assumed to be 0.14 ppm. For relamping programs, the discarded lamps are packed in the corrugated

containers from which the new lamps were taken and are then loaded into enclosed vans or trailers for

removal. In this case, fewer lamps are broken; a 10 percent breakage was assumed (Truesdale et al.,

1993).

The modeling exercise predicted that approximately 6 percent of the mercury being transported

by garbage trucks and from group re-lamping is emitted to the atmosphere. This amounts to 1.4 Mg/year

(1.5 tons/year).

Mercury emissions from transporting and storing lamps sent to recycling plants were also

estimated using the waste pile emission model. Emissions were based on a 30-day storage time and an

average of 5 percent breakage for the transport and storage steps. Emissions from storage facilities were

estimated to comprise about 90 percent of the recycling transport and storage emissions, amounting to

approximately 0.008 Mg. Total mercury emissions from transport and storage of waste lamps is

estimated to be 0.01 Mg, or 0.04 percent of the mercury from lamps entering the MSW (Truesdale et al.,

1993) or 1.4 Mg/year (1.5 tons/year) total from lamp breakage during transport and storage.

The industry estimate of mercury emissions from discarded fluorescent lamps is 0.16 Mg/year

(0.18 tons/year) (National Electrical Manufacturers Association, 1995). The industry estimate assumes

that most lamps are land filled within a couple of days after their disposal and are covered with 0.5 to 1
3-7

foot of soil at that time. Simulating this land filling practice and measuring the amount of mercury

released led to an estimated mercury evaporation rate of 0.8 percent after 20 days when the lamps were

covered by 0.5 feet of soil, and 0.2 percent after 20 days when the lamps were covered by 1 foot of soil

(rather than the 6.6 percent estimated in Truesdale et al., 1993, which is the basis for U.S. EPA's

estimate). The 0.8 percent evaporation rate was used to calculate the annual rate of 0.16 Mg/year (0.18

tons/year). The National Electrical Manufacturers Association study also measured the maximum

mercury evaporation rate from a broken lamp to be 6.35 percent after 50 days. However, as explained

above, the industry calculation of national emissions assumes that all discarded lamps are covered by soil

within a couple of days of being discarded.

3.2 General Laboratory Use

Mercury is used in laboratories in instruments, as a reagent, and as a catalyst. In 1994, an

estimated 1.0 Mg (1.1 tons) of mercury were emitted into the atmosphere from general laboratory use. 

An emission factor of 40 kg of mercury emitted for each megagram of mercury used in laboratories was

estimated in a 1973 report (Anderson, 1973). Because this emission factor was based on engineering

judgment and not on actual test data, and because it is dated, the reliability of this emission factor is

questionable. From 1990 to 1992, there was a decline in mercury consumption in general laboratory use,

with consumption dropping from 32 Mg (35 tons) in 1990 to 18 Mg (20 tons) in 1992 (Bureau of Mines,

1992). However, the trend most recently has been slightly increasing consumption, with 24 Mg (26 tons)

in 1994 (Plachy, 1996) The annual emission estimate is the product of this consumption rate and the

emission factor noted above. The limitations of that emission factor make the emission estimate

uncertain.

3.3 Dental Preparation and Use

Mercury is used in the dental industry, primarily in amalgam fillings for teeth, although it may

also be used in other dental equipment and supplies. In 1995, an estimated 0.64 Mg (0.7 ton) of mercury

was emitted from dental preparation and use. This is an underestimate because it is derived using an

emission factor that applies only to emissions of mercury from spills and scrap during dental preparation

and use (2 percent of mercury used is emitted into the atmosphere) (Perwak, 1981). The total amount of

mercury used in the dental industry is 31 Mg (34 tons) and includes mercury used in all dental equipment

and supplies, not just the amount used in dental preparation and use (Plachy, 1997). Mercury air

emissions not accounted for in dental preparation and use are most likely accounted for in the emission

estimates for municipal waste combustors, medical waste incinerators, and crematories. Mercury

discharges from dental offices to publicly owned sewage treatment facilities are also known to occur but

are not addressed in this report.

3.4 Municipal Solid Waste Landfills

As discussed throughout this volume, a variety of mercury-containing wastes are disposed in

non-hazardous (municipal and industrial) and hazardous waste landfills. These landfills can serve as

broad sources of airborne emissions of mercury as the disposed materials are broken or degraded, not

only while the landfill is actively receiving and disposing of wastes but also after the land filling stops

and waste materials are covered with soil. 
3-8

Municipal solid waste (MSW) landfills are landfills used primarily for the disposal of nonhazardous household wastes. Mercury is emitted from MSW landfills as a trace constituent of landfill

gas, which may be produced through anaerobic decomposition of waste. Measurement data of mercury

emissions were obtained for selected landfills that range from 7.0 x 10 ppm to 2.5 x 10 ppm ESCOR, -7 -3

Inc., 1982; Myers, 1996). From these measurements, EPA has calculated an average mercury

concentration in landfill gas to be 2.9 x 10 ppm. By combining this value with the 1994 estimate of -4

total landfill gas emitted of 10.2 million Mg (11.2 million tons) (EPA, 1995c), total 1994 emissions of

mercury from MSW landfills have been estimated to be 0.074 Mg (0.081 tons). Note that this figure

does not include emissions from industrial and hazardous waste landfills.

3.5 Mobile Sources

Mobile sources are defined in this report as diesel- and gasoline-powered, on-road, light-duty

vehicles. Of these types, gasoline-powered vehicles make up the most significant mobile emission

sources. A 1983 study indicated an estimated mercury emission factor of 1.3 x 10 milligram per -3

kilometer (mg/km) (4.6 x 10 pound per mile [lb/mile]) traveled for tail-pipe emissions from motor -9

vehicles (Pierson and Brachaczek, 1983). These data were for particulate mercury emissions derived

from neutron activation analysis of particulate filters. The population of vehicles studied was

81.9 percent gasoline-powered passenger cars, 2.4 percent gasoline-powered trucks and 15.7 percent

diesel trucks. The data are of questionable reliability for the current vehicle population because this

emission factor is based on a 1977 ambient sampling study, which predated the broad use of catalytic

converters and unleaded gasoline, widely mandated `State-regulated inspection and maintenance

programs and diesel-powered vehicle emission control requirements. It is unknown what effect these

measures might have on mercury emissions.

A 1979 study characterized regulated and unregulated exhaust emissions from catalyst and noncatalyst equipped light-duty gasoline-powered automobiles operating under malfunction conditions

(Urban and Garbe, 1979). An analysis for mercury was included in the study, but no mercury was

detected in tail-pipe emissions. The analytical minimum detection limit was not stated. A 1989 study

measured the exhaust emission rates of selected toxic substances for two late model gasoline-powered

passenger cars (Warner-Selph and DeVita, 1989). The two vehicles were operated over the Federal Test

Procedure (FTP), the Highway Fuel Economy Test (HFET) and the New York City Cycle (NYCC). 

Mercury was among the group of metals analyzed but was not present in detectable quantities. The

analytical minimum detection limits for mercury in the three test procedures were the following: FTP

0.025 mg/km (8.9 x 10 lb/mile) HFET 0.019 mg/km (6.7 x 10 lb/mile) and NYCC 0.15 mg/km (53.2 x -8 -8

10 lb/mile) (Warner-Selph and Lapp, 1993). These minimum detection limits are more than ten times -8

higher than the estimated emission factor presented in the 1983 study.

Given the uncertainties associated with these data, tail-pipe mercury emissions from mobile

sources were not calculated. The U.S. EPA also recognizes that various components of motor vehicles

may contain mercury (e.g., certain truck and hood light switches, used motor oil, certain headlights and

remote controls). Mercury emissions from the disposal or breakage of these components were not

estimated in this study. The potential for mercury emissions from other types of mobile sources,

including ships, were not assessed in this study.

3.6 Paint Use

Four mercury compounds -- phenylmercuric acetate, 3-(chloromethoxy) propyl mercuric acetate,

di(phenyl mercury) dodecenylsuccinate, and phenylmercuric oleate -- have been registered as biocides
3-9

for interior and exterior paint (U.S. EPA, 1990). Mercury compounds are added to paints to preserve the

paint in the can by controlling microbial growth. Prior to 1991, much larger amounts of mercury were

added to preserve the paint film from mildew after paint was applied to a surface. During and after

application of paint, these mercury compounds can be emitted into the atmosphere. As of May 1991, all

registrations for mercury biocides used in paints were voluntarily canceled by the registrants, thus

causing a drastic decrease in the use of mercury in paint (Agocs et al., 1990). In addition to the paint

industry reformulating its paints to eliminate mercury, U.S. EPA banned the use of mercury in interior

paint in 1990 and in exterior paint in 1991. The paint industry's demand for mercury in 1989 was

192 Mg (211 tons) but fell to 6 Mg (7 tons) in 1991, and had been completely eliminated in 1992

(Bureau of Mines, 1992).

Because Bureau of Mines data show no mercury usage in paint in 1992, emissions from this

source were assumed to be zero. This presumes that all mercury emissions are generated from paint

application the year the paint was produced. The U.S. EPA recognizes that current stocks of paint that

are still being sold may include paint that contains mercury. Data were unavailable to estimate potential

mercury emissions from this existing paint supply.

Prior to 1992, latex paints contributed significantly to atmospheric emissions. A 1975 study,

performed when the demand for mercury in paint was high, estimated that 66 percent of the mercury

used in paints was emitted into the atmosphere (Van Horn, 1975). Limited information suggests that

emissions could occur for as long as seven years after initial application of paint to a surface, although

the distribution of emissions over this time period is unknown (U.S. EPA, 1992a). Even so, this source

category is a good example of past industrial uses and releases of mercury to the environment. 

Assuming the estimate is correct that 66 percent of the mercury in paint is emitted, as recently as 1989 as

many as 140 tons of mercury were emitted from paint application alone in one year. Whether current

levels of mercury in the environment are more likely the result of historical emissions like these or are

attributable to current anthropogenic sources is still being debated.

3.7 Agricultural Burning

Mercury contamination of freshwater fish in the Florida Everglades has led to the investigation

of possible mercury sources in south Florida. The preharvest burning of sugarcane has been proposed as

a potential source of mercury to this area. One study estimated the atmospheric loading of mercury from

burning sugarcane stalks and leaves and muck soils (Patrick, et al., 1994). An emission factor of 0.0002

kg mercury per hectare of burned crop was calculated. This resulted in 0.036 Mg (0.04 tons) of mercury

emitted to the atmosphere from the preharvest burning of 174,00 acres of the Everglades Agricultural

Area sugarcane crop.

Other types of agricultural burning may also contribute to mercury emissions, for example landclearing activities. For this report, a national estimate of mercury emissions from sugarcane burning or

other agricultural activities was not calculated because of the limited emissions data and a lack of data on

the magnitude and frequency of these activities. The above study is presented to illustrate the potential

magnitude of mercury from these activities in one area of the country.
3-10

3.8 Other Area Sources

Sludge application is another recognized area source of airborne emissions of mercury. This

includes the agricultural and lawn application of municipal sewage sludge, which contains a number of

nutrients beneficial to plants, as well as the land application of municipal and industrial sludges as a

disposal method. Insufficient data were available to estimate national emissions of mercury from this

activity.
4-1

Table 4-1

Best Point Estimates of Annual Mercury Emissions from Combustion, Manufacturing and

Miscellaneous Point Sources: 1994-1995

Source

Emissions

Mg/yr Tons/yr

Combustion 125.3 137.7

Manufacturing 14.4 15.6

Miscellaneous 1.3 1.4

4. ANTHROPOGENIC POINT SOURCES OF MERCURY EMISSIONS

A point source is a stationary location or fixed facility from which pollutants are discharged or

emitted. Point sources account for approximately 98 percent of mercury emissions from anthropogenic

sources. Table 4-1 presents the estimated aggregate mercury emissions from combustion, manufacturing

and miscellaneous point sources. The sections that follow discuss the basis for the point source estimates

for each source category within these three groups.

4.1 Combustion Sources

Combustion sources include utility boilers, medical waste incinerators, municipal waste

combustors, commercial/industrial boilers, hazardous waste combustors, residential boilers, wood

combustion, sewage sludge incinerators and crematories. Mercury emissions from these sources

(excluding wood-fired residential heaters) account for an estimated 125 Mg/yr (138 tons/yr) or 87

percent of the mercury emissions generated annually in the United States. These types of combustion

units are commonly found throughout the country and are not concentrated in any one geographic region. 

Information concerning emissions, fossil fuel consumption on a per-State basis and location is presented

in Appendix A. 

Mercury exists naturally as a trace element in fossil fuels and can also be found in wastes. It is a

highly volatile metal that vaporizes at the temperatures reached during the combustion zones of the

processes discussed here. Consequently, mercury is emitted as a trace contaminant in the gas exhaust

stream when waste materials containing mercury or fuels such as coal, oil, or wood are fired.

This section provides background information on each of the combustion sources and discusses

the methodology used to estimate mercury and mercury compound emissions from the following: 

(1) utility boilers, (2) municipal waste combustors (MWCs), (3) commercial/industrial boilers,

(4) medical waste incinerators (MWIs), (5) hazardous waste combustors, (6) residential boilers, (7)

sewage sludge incinerators (SSIs), (8) wood combustors, and (9) crematories. For each of these source

types, processes and control measures currently in place are discussed, along with emission estimates and

the bases for those estimates. When a high degree of uncertainty within specific data is known, it 

is noted. Table 4-2 presents the estimated emissions from each source category.
4-2

Table 4-2

Best Point Estimates of Mercury Emissions from Anthropogenic Combustion Point Sources: 1994

Emissions

Source Mg/yr Tons/yr % of total Date of Data Uncertainty Basis for Em

a

Degree of

b

Utility boilers 47.2 51.8 32.8 1990/1994 Medium Test data; industry (E - coal (46.9) (51.6) Institute) estimates ar - oil (0.2) (0.2) fired utilities - natural gas (0.002) (0.002)

c

Municipal waste combustors 26.9 29.6 18.7 1986-92/1991 Medium Test data eCommercial/Industrial boilers 25.8 28.4 17.9 /1994 High Mass balance; emissio

 - coal (18.8) (20.7) because emission fact - oil (7.0) (7.7)

d

Medical waste incinerators 14.6 16.0 10.1 1996/1996 Medium Test data eHazardous waste combustors 6.4 7.1 4.4 1996/1996 Medium Test dataResidential boilers 3.3 3.6 2.3 /1994 High Mass balance; emissio

 - coal (0.4) (0.5) because emission fact - oil (2.9) (3.2)

d

Sewage sludge incinerators 0.9 1.0 0.6 1995/1996 High Test dataWood-fired boilers 0.2 0.2 0.1 1984-92/1980 Medium Test data fCrematories 0.0005 0.0006 0.0 1992/1995 High Engineering judgmenTotal 125.2 137.9 86.9

 Date that data emission factor is based on/date of activity factor used to estimate emissions. a

 A "medium" degree of uncertainty means the emission estimate is believed to be accurate within + 25 percent. A "high" degree of uncertainty means the emission esti

b

within + 50 percent.

 Parentheses denote subtotal within a larger point source category. c

 Date of data used to develop emission factor was not determined. d

 EPA has finalized emissions guidelines for these source categories which will reduce mercury emissions by at least an additional 90 percent over 1995 levels. e

 Does not include residential wood combustion emissions. f
4-3

 4.1.1 Utility Boilers

Utility boilers are large boilers used by public and private utilities to generate electricity. Such

boilers can be fired by coal, oil, natural gas, or some combination of these fuels (U.S. EPA, 1993a). 

Figures 4-1 and 4-2 show the locations of operating coal-fired and oil-fired utility boilers across the

United States, respectively.

In 1994, utility boilers consumed fossil fuel at an annual level of 20 x 10 megajoules (MJ) 12

(21 x 10 British thermal units [Btu]). About 81 percent of this total energy consumption resulted from 15

coal combustion, 4 percent from oil and petroleum fuels and 15 percent from natural gas consumption

(U.S. Department of Energy, 1996). In terms of coal usage, the majority of total nationwide coal

combustion (about 86 percent) is in utility boilers. Almost all of the coal burned in the U.S. is

bituminous and subbituminous (95 percent) while only 4 percent is lignite (Brooks, 1989). The

combustion processes used for these different coals are comparable. The most common liquid fuel used

by utility boilers is fuel oil derived from crude petroleum. Fuel oils are classified as either distillate or

residual.

4.1.1.1 Description of the Different Utility Boiler Types

Because there is no evidence to show that mercury emissions are affected by boiler type, this

section presents only a brief discussion of different boiler types and combustion techniques. More

information on boiler types may be found in the Air Pollution Engineering Manual, AP-42, Steam: Its

Generation and Use, and the L&E document (Buonicore and Davis, 1992; U.S. EPA, 1988a; Babcock

and Wilcox, 1975; U.S. EPA, 1997a).

Although several options are available for each component of a utility operation, the overall

process for coal-fired utility boilers is straightforward. Coal is received at the plant, typically by rail or

barge, unloaded and transferred to storage piles or silos. From storage, the coal is subjected to

mechanical sizing operations and then charged to the boiler. Coal-fired boilers are typically suspensionfired pulverized coal or cyclone systems. The other major process component is the ash-handling system

for the bottom ash and the fly ash that is collected in the air pollution control system (U.S. EPA, 1988a).

Oil-fired utility boilers are even simpler and have less variation in design than do the coal-fired

systems. Oil is received by barge, rail, truck, or pipeline and transferred to storage tanks. From there the

oil is fired to the boiler system. The main components of the system are the burner and the furnace. The

primary difference in systems that fire distillate and residual oils is the presence of an oil preheater in

residual systems (U.S. EPA, 1988a; Buonicore and Davis, 1992).


4-6

4.1.1.2 Effectiveness of Particulate Matter and Acid Gas Air Pollution Controls for Mercury

Although small quantities of mercury may be emitted as fugitive particulate matter (PM) from

coal storage and handling, the primary source of mercury from both coal and combustion in utility

boilers is the combustion stack. Because the combustion zone in boilers operates at temperatures above

1100C (2000F), mercury in the coal and oil is vaporized and exhausted as a gas. Some of the gas may

cool and condense as it passes through the boiler and the air pollution control device (APCD). The

primary types of control devices used for coal-fired utility boilers include electrostatic precipitators

(ESPs); wet scrubbers; fabric filters or baghouses (FFs), which are typically used as a component of a

dry flue gas desulfurization system (FGDs); and mechanical collectors. Mercury control efficiencies for

each of the control devices are presented in Figure 4-3. The test data used to calculate the removal

efficiencies described below are shown in more detail in Appendix B.

ESPs are the most widely used control device by the fossil fuel-fired electric utility industry. 

Because mercury in electric utility flue gas is predominantly in the vapor phase (Clarke and Sloss, 1992),

with only about 5 to 15 percent in the fly ash (Noblett et al., 1993), ESPs are relatively ineffective at

removing mercury compounds from flue gases. Cold-side ESPs, located after the air preheater have a

median mercury removal efficiency of 14.7 percent for coal-fired units, with actual test data ranging

from no control (zero percent removed) to 82.4 percent removal (Interpoll Laboratories, 1992a; Interpoll

Laboratories, 1992b; Interpoll Laboratories, 1992c; Radian Corporation, 1993a; Interpoll Laboratories,

1992d; Interpoll Laboratories, 1992e; Radian Corp., 1992a; Radian Corp., 1993a; Radian Corp., 1993b;

Radian Corp., 1993e; Radian Corp., 1994a; Battelle, 1993a; Battelle, 1993c; EPRI, 1993a; EPRI, 1993b;

EERC, 1993; Weston, 1993b; and Southern Research Institute, 1995a). Cold-side ESPs were found to

have a median mercury removal efficiency of about 62.4 percent in two tests of oil-fired units, with a

range from 41.7 to 83 percent removal (Carnot, 1994b; Carnot, 1994c). Data from one emission test for

a hot-side ESP, located before the air preheater, indicated no mercury control on a coal-fired unit

(Southern Research Institute, 1993b).

Scrubbers or FGD units for coal-fired plants are generally used as devices for removal of acid

gases (mainly SO emissions). Most utility boilers have an ESP or a FF before the wet FGD units to 2

collect the majority of PM. FGD units have a median mercury removal efficiency of about 22.6 percent,

with a range from 0 percent to 61.7 percent removal (Interpoll Laboratories, 1991; Interpoll Laboratories,

1990a; Radian Corp, 1993a; Radian Corp, 1993b; Radian Corp, 1994b; Radian Corp, 1994c; Radian

Corp, 1994d; EPRI, 1993a, Battelle, 1993a). One emission test across an ESP/wet-FGD (spray-tower

absorber) system showed a mercury removal efficiency of 82 percent (Radian Corporation, 1993b).

A spray dryer adsorber (SDA) is a dry scrubbing system followed by a particulate control device. 

A lime/water slurry is sprayed into the flue gas stream and the resulting dried solids are collected by an

ESP or a FF. Tests conducted on a SDA/FF system had a median mercury removal efficiency of

24 percent, with a range from 0 percent to 55 percent removal (Radian 1993c; Southern Research

Institute, 1993a; Interpoll Laboratories, 1991; Interpoll Laboratories, 1990b).
100

90

80

70

60

50

40

30

20

10

0

FGD(6) SDA(4) Fabric Filter(5) Cold-Side ESP/

Coal Fired(17)

Hot-Side ESP/

Coal Fired(2)

Cold-Side ESP/

Oil Fired(2)

Minimum

Maximum

Mean

Numbers in parentheses are numbers of test results.

Bars represent the standard deviation around the mean.

Data and references used to produce this figure are presented in Appendix B.

Figure 4-3

Comparison of Mercury Efficiencies Without Activated Carbon Injection

4-7
4-8

Fabric filters are more effective than ESPs at collecting fine particles. This performance may be

important in achieving better mercury removal. Also, the mercury may adsorb onto the fly ash cake that

is collected on the fabric and allow more residence time for mercury removal. FFs have a median

mercury removal efficiency of 8 percent, with a range from no control (zero percent removal) to

73 percent removal (Radian Corporation, 1993d; Carnot, 1994a; Interpoll, 1992d; Battelle, 1993b;

Weston, 1993a).

Mechanical collectors typically have very low PM collection efficiencies, often lower than

20 percent for particles less than or equal to 1 μm in size. These devices are used as gross particulate

removal devices before ESPs or as APCDs on oil-fired units. Venturi scrubbers can be effective for

particulate control, but require high pressure drops (more than 50 or 60 in. of water) for small particles. 

Even with high pressure drops, ESPs and FFs are normally more effective for submicron particles. 

Mechanical collectors and venturi scrubbers are not expected to provide effective mercury removal,

especially for those mercury compounds concentrated in the sub-micron PM fractions and in the vapor

phase.

4.1.1.3 Estimated National Mercury Emissions from Utility Boilers

To estimate national mercury emissions from utility boilers, data were gathered on the type of

fuel burned, the mercury content of each fuel and the amount of fuel consumed per year by each

individual unit (boiler). Data on plant configurations, unit fuel usage and stack parameters (on a boilerspecific basis) were obtained from the Utility Data Institute (UDI)/Edison Electric Institute (EEI) Power

Statistics database (1995 edition). The UDI/EEI database is compiled from Form EIA-767, which

electric utilities submit on a yearly basis to the U.S Department of Energy's Energy Information

Administration. Emissions were only calculated for operational or stand-by units. Previous estimates

were based on the assumption that all the mercury present in the fuel would be emitted in the stack gas

(U.S. EPA, 1993d). In addition, previous estimates did not attribute any mercury reductions to coal

cleaning. As explained below, the estimates presented in this report do account for reductions in the

mercury content of coal due to coal cleaning and considers any mercury reductions achieved by existing

control devices.

Calculation of utility mercury emissions was a two-step process. First, the amount of mercury in

the fuel was estimated as described below. The calculated mercury concentration in the fuel multiplied

by the fuel feed rate resulted in an estimate of the amount of mercury (in kg/year) entering each boiler. 

Next, based on test data, "emission modification factors" (EMFs) were developed that are specific to

various boiler configurations and control devices. The EMFs basically represent the level of mercury

control seen across various boiler configurations and control devices. (The control devices are those that

are currently installed on boilers principally for nitrogen oxide, sulfur dioxide and PM control.) The

EMFs developed from the tested units were applied to all other similar units in the U.S. to give mercury

emission estimates on a per-unit basis.

Only coal, oil and natural gas were considered because these fuels account for nearly 100 percent

of the fuels fired by utility boilers. The mercury content of these fuels varies greatly, with coal

containing the most mercury and natural gas containing almost none.

Mercury Concentrations in Oil and Natural Gas. The mercury concentration in as-fired oil and

natural gas was estimated from emissions test data for boilers burning these fuels. In the estimation of

mercury emissions, all oil-fired units were assumed to burn residual oil because trace element data were

available only for residual oil. An average density of 8.2 lb/gal was chosen to represent all residual oils. 
4-9

Trace element analysis of natural gas was performed for only two available emissions tests; these

concentrations were averaged. The calculated mercury concentration in the oil and natural gas

multiplied by the fuel feed rate resulted in an estimate of the amount of mercury (in kg/year) entering

each oil- and natural gas-fired boiler.

Mercury Concentrations in Coal. Mercury concentrations were estimated for bituminous,

subbituminous and lignite coals. The mercury concentration of anthracite coal was not calculated

because only 6 (out of approximately 2000) utility boilers fire anthracite and account for only 0.4 percent

of the coal burned annually. For the purposes of calculating mercury emissions, units burning anthracite

were assumed to burn bituminous coal.

A database of trace element concentrations in coal, by state of coal origin, was compiled by the

United States Geological Survey (USGS), which analyzed 3,331 core and channel samples of coal. 

These samples came from 50 coal beds having the highest coal production in the U.S. Industry reviewed

these data and under a separate effort screened the data to remove about 600 entries representing coal

seams that could not be mined economically (EPRI, 1994). The mercury concentration of the screened

data set was virtually the same as the mercury concentration when the full USGS data set was used, so

U.S. EPA chose to use the USGS data in its entirety. The mercury concentration of the samples ranged

from 0.003 ppmwt to 3.8 ppmwt (Bragg, 1992).

The average mercury content of each of these beds was calculated. The location of each bed was

then matched with a state. Using the UDI database and records of actual coal receipts, the state from

which each utility purchased the majority of its coal was identified. With three exceptions, the mercury

content of the coal fired by each utility was then assigned based on the average concentration of mercury

calculated for each coal bed. Exceptions were made for Colorado bituminous, Illinois coal, and

Wyoming coal where data were available from as-fired coal samples. These data were used directly to

estimate emissions from utility boilers firing these coals. There were two sets of data for coal originating

in Arizona and Washington. These two sets were averaged. Since no data were avail-able for coal from

Louisiana, data from Texas lignite coal were substituted for Louisiana lignite coal.

Mercury Reductions Due to Coal Cleaning. The USGS database contains concentrations of

mercury in as-mined coal but does not include analyses of coal shipments (i.e., "as-fired" coal). The

concentration of mercury in as-mined coal may be higher than the concentration in shipped coal because

in the process of preparing a coal shipment, some of the mineral matter in coal - and the associated

mercury - may be removed by coal cleaning processes. Since approximately 77 percent of the eastern

and midwestern bituminous coal shipments are cleaned in order to meet customer specifications for

heating value (Akers et al., 1993), ash and sulfur content, analyses were done to estimate the average

amount of mercury reduction that could be attributed to coal cleaning. As a result of these analyses, a 21

percent reduction in mercury concentration was attributed to coal cleaning for those boilers purchasing

coal from states where coal washing is common practice. The highlight box below discusses how this

mercury reduction value was determined. No coal cleaning reductions were applied to lignite or

subbituminous coals, or bituminous coal when the state of coal origin was west of the Mississippi River.
4-10

EFFECT OF COAL CLEANING ON MERCURY CONCENTRATIONS

U.S. EPA requested data on the concentrations of trace elements (including mercury) in coal from

the National Coal Association, but limited data were available for two reasons. First, few shipments are

analyzed for trace element concentrations, and second, many coal companies consider such information

proprietary. EPA did receive data on the concentrations of trace elements in coal shipments from the ARCO

Coal Company on 145 samples of Wyoming coal and on 30 samples of bituminous Colorado coal; the Illinois

State Geological Survey (ISGS) on 34 samples of Illinois coal; and the Electric Power Research Institute

(EPRI) on mercury concentrations in 100 various samples.

Since no other data were available on the concentration of mercury in actual coal shipments,

arithmetic averages of the mercury concentrations provided by the ARCO Coal Company and the ISGS were

considered as-fired samples. These values were used directly to estimate the amount of mercury in

bituminous Colorado coal, subbituminous Wyoming, and bituminous Illinois coal shipments. 

The mercury concentrations in the raw coal, the clean coal, and the percent reduction achieved by

cleaning are shown in Table 4-3. As shown, some of the mercury reductions are negative. At first, this

would seem to suggest that the mercury has been increased or enriched in the clean coal. Negative

percentages occur when part of the coal is removed, but the mercury is not contained in the extracted portion. 

As a result, the same weight of mercury that was contained in the uncleaned coal is contained within a

relatively smaller weight of the cleaned coal. Since the weight of the mercury was not changed, negative

removal percentages were interpreted to mean that no mercury reduction occurred, or in other words, that the

mercury reduction was zero percent.

As shown in Table 4-3, the mercury reductions ranged from -200 percent (effectively zero percent

removal) to 64 percent. There is also variation in mercury reduction from cleaned coals originating from the

same coal seam. For example, the mercury reduction ranged from -20 percent to 36 percent for Pittsburgh

seam coals. The variation may be explained by several factors. The data may represent different cleaning

techniques, and the effectiveness of the cleaning processes will depend on how much mercury was contained

in the coal. Also, considerable variation may result from the mercury analytical technique.

Because of the variability of the data, typical mercury removal was estimated by taking the

arithmetic average of the removal data listed in Table 4-3. Any negative value was taken as a zero, and the

zero values were included in the average. The resulting 21 percent average reduction was used to estimate

mercury emissions from utility boilers that burn bituminous coal from states east of the Mississippi River. 

Note that this reduction was assumed for all such boilers, even though data indicate that only 77 percent of

the eastern and midwestern bituminous coal shipments are cleaned. As stated above, no coal cleaning

reductions were applied to lignite or subbituminous coals, or bituminous coal when the state of coal origin

was west of the Mississippi River.

As these data demonstrate, coal cleaning can result in mercury reductions that are higher or lower

than the average 21 percent value applied in this analysis. It is expected that significantly higher mercury

reductions can be achieved with the application of emerging coal preparation processes, such as selective

agglomeration and advanced column floatation.
4-11

Table 4-3

Comparison of Mercury Concentrations in Raw and Cleaned Coal

Seam State Raw Coal

Mercury (ppm)

Cleaned Coal

Mercury (ppm)

Percent

Removal

Central Appalachian Coal Sample A 0.09 0.1 -11.11

Central Appalachian Coal Sample B 0.12 0.11 8.33

Il #6 IL 0.14 0.08 42.86

Pittsburgh A PA 0.15 0.11 26.67

Pittsburgh B PA 0.14 0.09 35.71

Pittsburgh C PA 0.14 0.13 7.14

Pittsburgh D PA 0.1 0.12 -20.00

Pittsburgh E PA 0.1 0.08 20.00

Pittsburgh PA 0.1 0.08 20.00

Upper Freeport PA 0.03 0.09 -200.00

Lower Kittanning PA 0.44 0.34 22.73

Sewickley PA 0.18 0.18 0.00

Pittsburgh PA 0.13 0.11 15.38

Pittsburgh PA 0.13 0.12 7.69

Il #6 IL 0.12 0.13 -8.33

KY #9 and 14 KY 0.16 0.14 12.50

Pratt/Utley AL 0.28 0.22 21.43

Pratt AL 0.29 0.28 3.45

Utley AL 0.34 0.27 20.59

Pratt AL 0.34 0.24 29.41

Upper Freeport PA 0.7 0.25 64.29

Upper Freeport PA 0.7 0.28 60.00

Il 2,3,5 IL 0.24 0.2 16.67

Il 2,3,5 IL 0.24 0.14 41.67

Ky #11 KY 0.15 0.12 20.00

ISGS IL 0.2 0.09 55

Minimum -200.00

Maximum 64.29

Average 21.21

Reference: Akers et al., 1993 for every seam but ISGS; Demir et al., 1993 for ISGS.
 Limestone is used in circulating fluidized bed (CFB) boilers to control sulfur dioxide emissions. The EPA 1

recognizes that the limestone may contribute to trace metal emissions, including mercury. For the 19 CFB units in

the U.S., the potential contribution of limestone to the unit's mercury emissions was included in the mercury

emissions estimate for each boiler. 

4-12

For example, for a unit burning bituminous coal, the amount of mercury entering the boiler was

estimated by multiplying the average mercury content of the coal (specific to state of coal origin) by 0.79

to account for a 21 percent reduction due to coal cleaning. This product was multiplied by the unit's

annual fuel consumption rate to give the inlet mercury in kg/year.

Calculation of Mercury Emission Estimates. Emissions data were available from 58 emission

tests conducted by U.S. EPA, the Electric Power Research Institute (EPRI), the Department of Energy

(DOE), and individual utilities. Not all known boiler configurations or control devices could be tested. 

In order to estimate emissions from all units in the U.S., EMFs were developed for specific boiler

configurations and control devices from the test data and applied to similar units.

The EMFs were calculated by dividing the amount of mercury exiting either the boiler or the

control device by the amount of mercury entering the boiler. The average EMF for specific boiler

configurations and control devices was calculated by taking the geometric mean of the EMFs for that

type of configuration or control device. (The geometric mean was chosen rather than the arithmetic

mean because the distribution of emission factors followed a lognormal distribution.) The EMFs for

various boiler configurations and control devices are shown in Appendix C. To calculate the control

efficiency, the EMF is subtracted from 1.

Boiler-specific emission estimates were then calculated by multiplying the calculated inlet

mercury concentration by the appropriate EMF for each boiler configuration and control device. 1

Figures 4-4 and 4-5 illustrate how mercury emission estimates were calculated for coal-fired boilers and

for oil- or natural gas-fired boilers. As displayed in Table 4-4, national estimates of mercury emissions

from utility boilers are approximately 52 tons per year, of which 51.6 tons are attributed to coal-fired

units, 0.2 tons are attributed to oil-fired units, and 0.002 tons are attributed to natural gas-fired units.

Table 4-4

Best Point Estimate of Mercury Emissions from Utility Boilers: 1994-1995

Fuel Type Mg/Yr Tons/Yr

Emission Rate Comments

Coal 46.9 51.6 The industry (Electric Power Research Institute)

estimate for coal-fired units is 44 tons/year.

Oil 0.2 0.2

Natural Gas 0.002 0.002

Total 47.2 51.8

PLANT

CONFIGURATION

INFORMATION

Subbituminous

Bituminous Lignite

USGS average coal

mercury concentration

specific to State of

coal origin

USGS average coal

mercury concentration

specific to State of

coal origin

Apply coal

cleaning factor,

if applicable No cleaning

factor

Mercury inlet to boiler

Multiply mercury content

of coal by unit annual feed

rate form UDI data base

What type of boiler?

Apply boiler trace element

emission factors (EMFs)

What type of particulate

matter (PM) control?

Apply PM trace element

emission factors

What type of SO2 control?

Apply SO trace element

emission factors

2

Kg/yr mercury out of stack

Identify State of coal

origin from UDI

Figure 4-5

Mercury Emissions from Coal Fired Boilers

4-14
4-15

4.1.2 Municipal Waste Combustors

Municipal waste combustors (MWCs) fire municipal solid waste (MSW) to reduce the volume of

the waste and produce energy. There are three main types of technologies used to combust MSW: mass

burn combustors, modular combustors and refuse-derived fuel-fired (RDF) combustors. A fourth type,

fluidized-bed combustors (FBCs), is less common and can be considered a subset of the RDF

technology. Modular MWCs characterize the low end of the MWC size range, whereas the mass burn

and RDF MWCs tend to be larger. Both the mass burn and modular MWCs fire waste that has

undergone minimal pre-processing, other than the removal of bulky items. The RDF combustors fire

MSW that has been processed to varying degrees, from simple removal of bulky and noncombustible

items, to extensive processing to produce a fuel suitable for co-firing in pulverized coal-fired boilers. Of

the three main combustor types, mass burn combustors are the predominant technology used and are

found in three kinds: mass burn/waterwall (MB/WW), mass burn/refractory wall (MB/REF) and mass

burn/rotary waterwall (MB/RC). The MB/WW technology is the most common type, especially for

newer MWCs. With the exception of the refractory wall combustors and some of the modular

combustors, the majority of MWCs incorporate energy recovery (Fenn and Nebel, 1992).

At the beginning of 1995, there were over 130 MWC plants with aggregate capacities greater

than 36 Mg/d (40 tons/d) of MSW operating in the United States. There have been a number of plant

closures in this source category since 1991. The inventory described here represents 37 fewer facilities

in this size range than reported by U.S. EPA in 1993 (U.S. EPA, 1993d). The number of combustion

units per facility ranges from one to six, with the average being two. Total facility capacity ranges from

36 to 2,700 Mg/d (40 to 3,000 tons/d). These plants have a total capacity of approximately 90,000 Mg/d

(99,000 tons/d). A geographic distribution of the MWCs is presented in Table A-8, Appendix A (Fenn

and Nebel, 1992). This distribution reflects MWC's that were operational in January 1995.

In addition to the MWCs discussed above, there are a number of smaller MWCs in the United

States (with plant capacities of less than 36 Mg/d [40 tons/d]). This population of smaller MWCs

comprises less than one percent of the nation's total MWC capacity (Fenn and Nebel, 1992). Since 1991,

there have been 13 MWCs in this size range that have closed. Table A-8 in Appendix A, as well as the

map shown in Figure 4-6, reflects the 1995 MWC population.

4.1.2.1 Mercury Emissions and Controls

Mercury emissions from MWCs occur when mercury in the MSW vaporizes during combustion

and is exhausted through the combustor stack. There are numerous sources of mercury in MSW. These

include electric switches and lighting components, paint residues and thermometers.

More than 85 percent of the MWC plants (99 percent of the MWC capacity) in the United States

employ some kind of APCD (Fenn and Nebel, 1992). These controls range from the use of electrostatic

precipitators (ESPs) alone to control PM, to the use of acid gas controls (e.g., dry lime injection, spray

drying) in combination with an ESP or a fabric filter. New MWCs employ the latter combination of

controls plus the application of activated carbon injection technology. Mercury control in APCDs

without supplemental carbon injection technology is variable since mercury exists as a vapor at the

typical APCD operating temperatures. Factors that enhance mercury control are low temperatures in the

APCD system (less than 150 to 200C [300 to 400F]), the presence of an effective
Figure 4-6

Municipal Waste Combustor Facilities

mercury sorbent and a method to collect the sorbent (Nebel and White, 1992). In general, carbon

present in the fly ash enhances mercury sorption onto PM, which can then be captured in the PM

control device. Most modem MWCs, excluding RDF combustors, have low levels of carbon in the fly

ash and good carbon burnout, representative of efficient and complete combustion; thus, there is little

carbon to adsorb the mercury. RDF combustors generally have higher PM loadings and higher carbon

contents at the combustor exit because of the suspension firing of the RDF in the combustor. As a

result, mercury levels for RDF MWCs with acid gas control alone (flue gas cooling) are lower than for

other combustors (Nebel and White, 1991). With the additional application of carbon injection

technology, non-RDF combustors achieve 85 to 95 percent mercury control with resulting emissions

similar to RDF combustors. Since 1994, 15 MWC units have initiated commercial operation with

carbon injection technology for mercury. The average performance level is 93 percent mercury

control.

Add-on mercury control techniques include the injection of activated carbon or Na2S into the

flue gas prior to the PM control system. These technologies are now being used commercially on some

MWCs in the U.S., and on MWCs in Europe, Canada and Japan where removal efftciencies have been

reported to range from over 50 percent to 90 percent. Recent test programs using activated carbon and

Na2S injection conducted in the U.S. showed mercury removal efficiencies ranging from 50 percent to

over 95 percent (U.S. EPA, 1993a). There are currently at least four MWCs in the U.S. that are being

controlled with activated carbon injection in conjunction with PM control. Greater than 95 percent

control of mercury emissions is being achieved. State regulations in Florida and New Jersey required

MWCs in these states to retrofit with activated carbon injection by the end of 1995.

4-16
4-17

New Source Performance Standards and

Emission Guidelines for MWCs

On September 20, 1994, the U.S. EPA proposed New Source Performance Standards (NSPS) and Emission

Guidelines (EG) applicable to MWC plants larger than 35 Mg/day (39 tons per day) capacity. The U.S. EPA finalized

these regulations on October 31, 1995. The NSPS (Subpart Eb) applies to new MWC plants constructed after September

20, 1994 and the EG (Subpart Cb) applies to MWC plants constructed before September 20, 1994. For some of the

pollutants regulated by the NSPS and EG, the NSPS is more stringent than the EG. For mercury, the same emission

control requirements apply to new MWCs (NSPS) and existing MWCs (EG). The final mercury standard for new and

existing MWCs is 0.08 mg/dscm or about 90 percent control.

Emission factors for mercury have been developed from test data gathered at several MWCs. 

The emission factors for various combinations of combustors and control devices are presented in

Table A-9, Appendix A. Estimated mercury emissions were determined based on the tonnage of the

waste being combusted and on these emission factors (U.S. EPA, 1992b; Waste Age, 1991). Multiplying

the processing rates by the uncontrolled emissions and taking into account the different control

efficiencies (all found in Table A-9, Appendix A) gives an estimated total baseline mercury emissions of

50 Mg/yr (55 tons/yr) in 1990. As described below, the 1995 emission estimate for MWCs is

considerably lower.

Mercury emissions from MWCs have declined since 1990 and will continue to decline in the

future for three important reasons. First, under section 129 of the CAA, U.S. EPA is required to develop

emission limits for mercury (and a number of other pollutants) being emitted from MWCs. On October

31, 1995, the U.S. EPA Administrator signed New Source Performance Standards (NSPS) and emission

guidelines for new and existing MWCs that have the capacity to burn more than 35 Mg MSW/day (39

tons/day) (see box below). The NSPS and emission guidelines, when fully implemented, are estimated

to reduce mercury emissions by about 90 percent, from the 1990 baseline of 50 Mg/year (55 tons/year) to

4.0 Mg/year (4.4 tons/year).

Second, as described in the following sections, many of the mercury-containing components that

comprise MSW have declined. These include household batteries where mercury use is expected to be

discontinued and paint residues and pigments where mercury additives have been phased out. Based on

the status of all MWC facilities in 1995, the U.S. EPA estimates national mercury emissions from

MWCs to be 26.9 Mg/yr (29.6 tons/yr). This estimate incorporates changes in MWC mercury emission

levels resulting from (1) installation of APCDs on new and some existing MWCs that achieve moderate

mercury control, (2) retirement of several existing MWCs, and (3) significant reductions in the mercury

content of mercury-containing components of municipal waste, as described above. As a result, the inlet

concentration of mercury in the MWC waste stream is estimated to be, on average, half of what the

concentration was in 1990. As mentioned above, full implementation of the 1995 emissions guidelines

(retrofit of carbon injection technology to existing MWCs) will result in national mercury emissions

from MWCs being reduced to 4.4 tons per year.

Third, some States have enacted either MWC legislation requiring the use of activated carbon

injection, recycling or bans on the sale of certain mercury-containing products. These efforts will

decrease both the amount of mercury being emitted from MWCs and the amount of mercury in MSW in

general. Florida, New Jersey and Minnesota have led State efforts in this area. Volume VIII of this

Mercury Report to Congress summarizes the legislative, regulatory and other programs of several states

that influence mercury use and disposal.
4-18

4.1.2.2 MSW Components and Trends

MSW consists primarily of household garbage and other commercial, institutional and industrial

solid wastes. The known sources of mercury in MSW are batteries (mercuric oxide), discarded electrical

equipment and wiring, fluorescent bulbs, paint residues and plastics. In 1989, the estimated mercury

content of MSW was 664 Mg (709 tons), with concentrations ranging from 1 to 6 ppm by weight and a

typical value being 4 ppm by weight (U.S. EPA, 1993a).

The U.S. EPA's Office of Solid Waste (OSW) estimates that 55 to 65 percent of MSW comes

from residential sources, while 35 to 45 percent comes from commercial sources (U.S. EPA, 1992g). 

One 1992 study identified and reported a number of specific sources of mercury in MSW, as summarized

in Table 4-5 The data from Table 4-5 are shown graphically for the year 1989 in Figure 4-7. These

figures show that in 1989 household batteries were the largest contributing source of mercury to MSW. 

Fluorescent light bulbs, paint residues, thermometers, thermostats, and pigments contribute most of the

remainder of mercury to MSW. However, as discussed in the subsections that follow, mercury in

batteries and paint residues have decreased significantly in the 1990s.

In general, from an examination of Bureau of Mines data for mercury use, it can be inferred that

the components of MSW that will be the main sources of mercury in the future will be in the electrical

lighting and wiring devices and switches sectors, as well as fever thermometers.

Batteries

Major types of batteries include alkaline, mercuric oxide, silver oxide, and zinc air batteries. 

Another kind of battery, carbon zinc, is produced and discarded at a substantially lower rate.

In 1989, alkaline batteries accounted for about 419 tons or close to 60 percent of the mercury in

the MSW stream (U.S. EPA, 1992a). Although the quantity of mercury in each alkaline battery is

minimal, the large number sold and discarded has made these batteries the largest single source of

mercury in MSW historically. The contribution from this source category, however, is declining

dramatically due largely to industry initiatives and recent federal and state laws to reduce and ultimately

eliminate mercury from alkaline batteries.

Mercury has been used in alkaline manganese batteries as an additive to suppress formation of

internal gases which would lead to leakage, possible explosions and/or short shelf life. In the U.S.,

alkaline batteries in the mid-1980's contained mercury in amounts from about 0.8 percent to about 1.2

percent of the battery weight. Between late 1989 and early 1991, all U.S. manufacturers converted

production so that the mercury content, except in button and "coin" cells, did not exceed 0.025 percent

mercury by weight (National Electrical Manufacturers Association, undated).
4-19

Table 4-5

Estimated Discards of Mercury in Products in Municipal Solid Waste a

In Tonsb,c

Products 1970 1975 1980 1985 1989 1995 2000

Batteries

Alkaline

Mercuric oxide

Others

Subtotal Batteries

4.1

301.9

4.8

310.8

38.4

287.8

4.7

330.9

158.2

266.8

4.5

429.5

352.3

235.2

4.5

592.0

419.4

196.6

5.2

621.2

*

*

*

*

0.0

*

0.0

*

Electric Lighting

Fluorescent Lamps

High Intensity Lamps

Subtotal Lighting

18.9

0.2

19.1

21.5

0.3

21.8

23.2

1.1

24.3

27.9

0.7

28.6

26.0

0.8

26.7

14.7d

1.0

15.7

11.6d

1.2

12.6

Paint Residues 30.2 37.3 26.7 31.4 18.2 2.3 0.5

Fever Thermometers 12.2 23.2 25.7 32.5 16.3 16.9 16.8

Thermostats 5.3 6.8 7.0 9.5 11.2 8.1 10.3

Pigments 32.3 27.5 23.0 25.2 10.0 3.0 1.5

Dental Uses 9.3 9.7 7.1 6.2 4.0 2.9 2.3

Special Paper Coating 0.1 0.6 1.2 1.8 1.0 0.0 0.0

Mercury Light Switches 0.4 0.4 0.4 0.4 0.4 1.9 1.9

Film Pack Batteries 2.1 2.3 2.6 2.8 0.0 0.0 0.0

Total Discards 421.8 460.5 547.5 730.4 709.0 227.6 144.6

 U.S. EPA, 1992a (except for fluorescent lamps estimates). a

 Discards before recovery. b

 One ton equals 2000 pounds. c

 The estimated contribution of mercury from fluorescent lamps disposal to MSW was calculated based on industry d

estimates of a 4 percent growth rate in sales in conjunction with a 53 percent decrease in mercury content between 1989

and 1995, and a further 34 percent decrease in mercury content by the year 2000 (to 15 mg of mercury per 4 foot

fluorescent lamp) (National Electric Manufacturers Association, 1995).

* NOTE: Since 1992 several states have restricted the mercury content of alkaline batteries and/or banned the sale of

mercuric oxide batteries. Federal legislation to restrict mercury use in batteries went into effect in May, 1996. The

battery industry has eliminated mercury as an intentional additive in alkaline batteries, except in button cells. Although

no current estimate of mercury emissions from batteries was available for these out years, according to NEMA, the entire

U.S. battery industry used only approximately 6.6 tons of mercury in 1994 (NEMA, 1996).
Figure 4-7

Discards of Mercury in Municipal Solid Waste, 1989

1.6%

2.3%

2.6%

Lighting 3.8%.

Total mercury discards = 709 tons

Mercuric oxide batteries include cylinder-shaped batteries (such as those used in hospital

applications) and button-shaped batteries (such as those used in hearing aids, electronic watches,

calculators, etc.). Larger mercuric oxide batteries are used in a variety of medical devices. The

mercury content of mercuric oxide batteries is 30 to 40 percent of the weight of the battery and cannot

be reduced without proportionately reducing the energy content of the battery. In 1989, these batteries

contributed an estimated 196 tons (or about 28 percent) of mercury discards to MSW. Although

mercuric oxide batteries are estimated to continue to be a large source of mercury in MSW on a

percentage basis (Solid Waste Association of North America, 1993), the total tonnage of mercury

discarded in such batteries is expected to decline in the future due to the increase in use of alkaline and

zinc air batteries for these applications. The Mercury-Containing and Rechargeable Battery

Management Act prohibits disposal of these batteries in the MWS after May 13, 1996 (see discussion

below).

Silver oxide, zinc air and carbon zinc batteries contributed an estimated 5 tons (or about

1 percent) of mercury discards in MSW in 1989. Because production of carbon zinc batteries is

declining, and because these batteries have been converted to “no mercury added” designs, discards of

mercury in carbon zinc batteries will decline. Production and discards of silver oxide and zinc air

batteries are increasing, but mercury use has been discontinued in these types of batteries since 1992

(National Electric Manufacturers Association, undated).

4-20
4-21

Table 4-6 presents the estimated amount of mercury entering the MSW stream by year and

battery type. However, it is important to note the estimates for the years 1995 and 2000 do not reflect

recent state, federal or battery manufacturers' efforts to reduce the mercury content of batteries.

A federal law called the Mercury-Containing and Rechargeable Battery Management Act went

into effect May 13, 1996. Under Title I: Rechargeable Battery Recycling Act, persons are prohibited

from selling for use in the United States a regulated battery (a rechargeable battery containing cadmium

or lead electrode, or other electrode chemistries determined by EPA) unless labeling requirements are

met and the battery is removable. The label must state that the battery must be recycled or disposed of

properly. Title II: Mercury-Containing Battery Act, prohibits the sale of 1) alkaline-manganese batteries

containing mercury (alkaline-manganese button cell batteries are limited to 25 mg mercury per button

cell), 2) zinc carbon batteries containing mercury, 3) button cell mercuric-oxide batteries for use in the

US, and 4) any mercuric-oxide battery unless the manufacturer identifies a collection site that has all

requires federal, State, and local government approvals, to which persons may send batteries for

recycling and disposal.

Several states has already passed or introduced legislation with similar requirements to the

federal law discussed above prior to the federal law’s effective date. With these restrictions on the

production and disposal of mercury containing batteries in MSW, mercury introduced into the waste

stream is expected to decrease over time.

The National Electrical Manufacturers Association (NEMA) has estimated that the average

mercury level in MSW from batteries will decline by 50% every two years and will be mercury freeby

approximately 2008 (NEMA, 1997). NEMA cautions readers that this projection of future mercury

levels is based on very few data and NEMA intends to conduct annual analyses to document the

continued decline in mercury levels. This estimate is based on results of the three analyses of samples of

post consumer round cell, alkaline manganese and zinc carbon batteries in the MWS. These were from

Camden County New Jersey battery drop off and collection program, the Lee County Florida battery

curbside collection program and the Hennepin County Minnesota drop off and curbside collection

programs. The study found that the most frequent (median) ages of alkaline batteries found in the

stockpile was 1-2 years old.

Electric Lighting

Fluorescent lamps (bulbs) and high intensity lamps (bulbs) used in lighting streets, parking lots,

etc. were considered the second largest source of mercury in MSW in 1989 (U.S. EPA, 1992a). It is

estimated that fluorescent lamps accounted for about 26 tons of mercury in MSW (or 3.7 percent of total

discards) in 1989. All lighting sources were estimated to contribute about 27 tons of mercury in the

same year. Figure 4-8 illustrates the estimated historical discards of electric lighting sources.

As indicated in the flow diagram in Figure 3-1, an estimated 98% of discarded bulbs are treated

as MSW (2% is estimated to be recycled). Of the bulbs in the MSW system, 13% are sent on to MWCs

for incineration. Approximately 90% of the mercury contained in these lamps would be expected to

volatilize and become emissions if there were no control device (Truesdale, 1993).
4-22

Table 4-6

Estimated Discards of Mercury in Batteriesa

In Tons

Alkaline Mercuric Oxide Silver Oxide Zinc Air Year Discarded

4.1 301.9 0.1 0.0 1970

38.4 287.8 0.2 0.2 1975

158.2 266.8 0.3 0.3 1980

352.3 235.2 0.5 0.7 1985

443.6 182.5 1.1 2.4 1990

390.5 172.0 1.1 2.9 1991

* * 0.7 2.0 1995

0.0 * 0.0 0.0 2000

 U.S. EPA, 1992a. a

* NOTE: Since 1992 several states have restricted the mercury content of alkaline batteries and/or banned the sale of

mercuric oxide batteries. Federal legislation to restrict mercury use in batteries went into effect in May, 1996. The

battery industry has eliminated mercury as an intentional additive in alkaline batteries, except in button cells. Although

no current estimate of mercury emissions from batteries was available for these out years, according to NEMA, the entire

U.S. battery industry used only approximately 6.6 tons of mercury in 1994 (NEMA, 1996).

As discussed in Section 3.1, EPA has proposed a new rule addressing the management of spent

mercury-containing lamps (59 FR 39288). One of the options considered in this proposal would be to

add mercury-containing lamps to the universal waste regulations, which would change the requirements

for lamp transport for recycling purposes. The second option would allow disposal of lamps in a Subtitle

D landfill, but would not allow the disposal of lamps in a MWC.

Future projections of mercury discards from electric lighting sources depend on the sales of

lamps and their mercury content. Sales of fluorescent lamps increase between 3 and 5 percent a year. As

described in section 3.1 of this Volume, the mercury content of fluorescent lamps has decreased by 53

percent between 1989 and 1995 to 22.8 mg of mercury per lamp. Assuming a 4 percent increase in sales

and a 53 percent decrease in mercury, estimated discards of mercury would be 14.7 tons in 1995. 

Assuming a 4 percent increase in sales and an additional 34 percent decrease in mercury content between

1995 and 2000 (to 15 mg mercury per lamp) leads to an estimated 11.6 tons per year in discards in the

year 2000.

4-24

Table 4-7

Estimated Discards of Mercury in Paint Residues a

Year Total Discards in Residues (In Tons)

1970 30.2

1975 37.3

1980 26.7

1985 31.4

1988 23.1

1990 17.5

1995 2.3

2000 0.5

 U.S. EPA, 1992a. a

Fever Thermometers

An estimated 16.3 tons of mercury were discarded in thermometers in 1989. It is estimated that

digital thermometers will gain an additional 1 to 2 percent of the market each year from 1990 through

2000, and the mercury content of mercury thermometers will remain constant (U.S. EPA, 1992a). Table

4-5 illustrates the estimated discards of mercury from thermometers in MSW from 1970 to 2000.

Thermostats

Mercury thermostats are being replaced with digital thermostats. It is expected that thermostats,

however, will still be a source of mercury in MSW through the year 2000 because of the long life of

mercury thermostats. Mercury thermostats contributed an estimated 11 tons of mercury to the MSW

stream in 1989 (U.S. EPA, 1992a). The estimated historical trends in mercury thermostat discards are

presented in Table 4-8. Federal legislation (the Universal Waste rule) finalized in 1995 encourages the

recycling of thermostats rather than their disposal. Recycling efforts are discussed in section 4.2.6.1 of

this Volume. As a result of recycling programs, mercury discards from thermostats are expected to

decline.
4-25

Table 4-8

Estimated Discards of Mercury in Thermostats a

Year Total Mercury (In Tons)

1970 5.3

1975 6.8

1980 7.0

1985 9.5

1988 10.7

1989 11.2

1995 8.1

2000 10.3

 U.S. EPA, 1992a. a

Pigments

Based on available data, one report estimated that 10 tons of mercury in pigments were discarded

in 1989. This accounted for less than 2 percent of total mercury discards. Most of the mercury used in

pigments is used in plastics, paints, rubber, printing inks, and textiles. As shown in Figure 4-9, estimated

discards of mercury in MSW pigments have generally been trending downwards since 1970 (U.S. EPA,

1992a).

Other MSW

Dental amalgams, a special paper coating used with cathode ray tubes, and mercury light

switches contributed less than 1 percent of the mercury in MSW in 1989. Plans are underway to

discontinue manufacture of the special paper by 1995. Mercury light switches are an increasing source

of mercury in MSW. One study projects that 2 tons of mercury will be discarded to MSW from mercury

light switches in the year 2000, which would account for about 1 percent of total discards in that year

(U.S. EPA, 1992a).

Several additional sources of mercury have been found in MSW, but have not been quantified. 

For example, mercury was a component of batteries used in instant camera film packs, but these batteries

were discontinued in 1988. Mirrors, glass, felt, outdoor textiles, and paper are other sources of mercury

to MSW.

4-27

As with utility boilers, the configuration of commercial/industrial boilers can vary, but the

overall system is straightforward. Coal or oil is received and transferred to storage where it is held until

it is transferred to the boiler. Because this source category encompasses a wide range of boiler sizes, the

types of boilers used are more varied than those used in the utility sector. Larger coal-fired industrial

boilers are suspension-fired systems like those used in the utility sector, while moderate and smaller

units are grate-fired systems that include spreader stokers, overfeed traveling and vibrating grate stokers

and underfeed stokers. Oil-fired furnaces, which may use either distillate or residual fuel oil, typically

comprise a burner, a combustion air supply system, and a combustion chamber. All coal-fired facilities,

and some oil-fired facilities, also have ash-handling systems.

Mercury emission factors for coal combustion in commercial/industrial boilers were developed

using mass-balance calculations with the assumption that all mercury fired with the coal is emitted in the

stack gas as a function of coal type (U.S. EPA, 1997a). The emission factors do not account for coal

washing because the U.S. EPA believes that buyers for commercial/industrial boilers do not purchase

washed coal; their source of coal is primarily the spot market. An estimated emission factor of

7.0 kg/10 J (16 lb/10 Btu) was used for bituminous coal combustion, and 7.6 kg/10 J (18 lb/10 Btu) 15 12 15 12

was used for anthracite coal combustion. Estimates of mercury emissions on a per- state basis from

coal-fired commercial/industrial boilers are provided in Table A-3, Appendix A. These values were

determined by using the referenced emission factors and the coal consumption estimates for the states

presented in Table A-2, Appendix A. In estimating emissions, it was assumed that mercury emissions

from commercial/industrial boilers were not controlled. The total estimated annual emissions for

coal-fired boilers are 18.8 Mg/yr (20.7 tons/yr). Because mercury reductions from coal washing and any

other reductions that may occur across existing control devices are not accounted for, the emissions may

be overestimated.

Mercury emissions for oil combustion in commercial/industrial boilers were estimated on a perstate basis using an emission factor of 2.9 kg/10 J (6.8 lb/10 Btu) for residual oil and 3.0 kg/10 J 15 12 15

(7.2 lb/10 Btu) for distillate oil and the oil consumption estimates for States given in Table A-2, 12

Appendix A. These calculated emission values are presented in Table A-4, Appendix A. The total

estimated annual emissions for oil-fired commercial/industrial boilers are 7 Mg/yr (7.7 tons/yr).

4.1.4 Medical Waste Incinerators

Medical waste incinerators (MWIs) are small incineration units that charge from 0.9 Mg/day

(1 ton/day) to 55 Mg/day (60 tons/day) of infectious and noninfectious wastes generated from facilities

involved in medical or veterinary care or research activities. These facilities include hospitals, medical

clinics, offices of doctors and dentists, veterinary clinics, nursing homes, medical laboratories, medical

and veterinary schools and research units, and funeral homes. The Resource Conservation and Recovery

Act (RCRA) (as amended November 1, 1988) defines medical waste as "...any solid waste which is

generated in the diagnosis, treatment, or immunization of human beings or animals, in research

pertaining thereto, or in the production or testing of biologicals" (U.S. EPA 1994a).

The estimated annual uncontrolled mercury emissions from MWIs are currently 14.6 Mg/yr

(16.0 tons/yr). In addition, the NSPS and emission guidelines for MWIs would decrease national

mercury emissions from MWIs by 94 percent, to an estimated level of 0.95 Mg/yr (1.0 ton/year) after

control (see the box below for more detail).

Several states including New York, California and Texas have adopted relatively stringent

regulations in the past few years limiting emissions from MWIs. The implementation of these
4-28

regulations has brought about very large reductions in MWI emissions of mercury in those states. It has

also significantly reshaped how medical waste is managed in those states. Many facilities have

responded to state regulations by switching to other medical waste treatment and disposal options to

avoid the cost of add-on pollution control equipment. The two most commonly chosen alternatives have

been off-site contract disposal in larger commercial incinerators and on-site treatment by other means

(e.g., steam autoclaving).

Mercury emissions from MWIs occur when mercury, which exists as a contaminant in the

medical waste, is combusted at high temperatures, vaporizes and exits the combustion gas exhaust stack. 

Known mercury sources in medical waste include batteries, fluorescent lamps, high-intensity discharge

lamps, thermometers, paper and film coatings, plastic pigments, antiseptics, diuretics, skin preparations,

pigments in red infectious waste bags and CAT scan paper. Much of the mercury in the medical waste

stream is thought to be emitted as mercuric chloride, due to the large amount of chlorinated plastic

products disposed.

U.S. EPA estimates that about 0.204 x 10 Mg/yr (0.268 x 10 tons/yr) of pathological waste and 6 6

1.431 x 10 Mg/yr (1.574 x 10 tons/yr) of general medical waste are processed annually in the United 6 6

States (U.S. EPA, 1993a). Medical waste may consist of any of the following, in any combination: 

human and animal anatomical parts and/or tissue; sharps (syringes, needles, vials, etc.); fabrics (gauze,

bandages, etc.); plastics (trash bags, IV bags, etc.); paper (disposable gowns, sheets, etc.); and waste

chemicals.

About 2,400 MWIs currently operate throughout the country; geographic distribution is

relatively even (see Table A-10, Appendix A) (U.S. EPA, 1996a). Most of these units are hospital

incinerators.

There are an additional 1,305 incinerators burning only pathological waste which are not

technically considered to be MWIs. These units are used for disposal of tissue only and are most

commonly found at veterinary facilities or animal research facilities. The primary source of mercury in

medical waste is mercury-containing products, not tissue. These small incinerators are estimated to

contribute 0.12 Mg/year (0.13 tons/year) to the total MWI mercury estimate of 14.6 Mg/year (16.0

tons/year). The reader should note that the NSPS and emission guidelines for MWIs do not apply to

either incinerators for pathological waste only or crematories. In this document, crematories are

discussed in Section 4.1.9.

The primary functions of MWIs are to render the waste biologically innocuous and to reduce the

volume and mass of solids that must be land filled by combusting the organic material contained within

the waste. Currently, three major MWI types operate in the United States: continuous-duty,

intermittent-duty and batch type. All three have two chambers that operate on a similar principle. Waste

is fed to a primary chamber, where it is heated and volatilized. The volatiles and combustion gases are

then sent to a secondary chamber, where combustion of the volatiles is completed by adding air and heat. 

All mercury in the waste is assumed to be volatilized during the combustion process and emitted with the

combustion stack gases.

 A number of air pollution control systems are used to control PM and gas emissions from MWI

combustion stacks. Most of these systems fall into the general classes of either wet or dry systems. Wet

systems typically comprise a wet scrubber, designed for PM control (venturi scrubber or rotary

atomizing scrubber), in series with a packed-bed scrubber for acid gas removal and a high-efficiency

mist elimination system. Most dry systems use a fabric filter for PM removal, but ESPs 
4-29

New Source Performance Standards and

Emission Guidelines for MWIs

On September 15, 1997, EPA finalized the NSPS for new MWIs and emission guidelines for existing MWIs

(62 FR 48348). The NSPS applies to all facilities that commenced construction after June 20, 1996 or that commenced

modification after the effective date of the NSPS (March 15, 1998), and the emission guidelines apply to existing MWIs

that commenced construction on or before June 20, 1996, although sources combusting only pathological wastes would

be subject to only certain reporting and record keeping provisions. Overall, the NSPS and emission guidelines implement

sections 111 and 129 of the Clean Air Act Amendments of 1990, including the requirement for MWIs to control

emissions of air pollutants to levels that reflect the maximum degree of emissions reduction achievable, taking into

consideration costs, any non-air-quality health and environmental impacts, and energy requirements (a standard

commonly referred to as "maximum achievable control technology" or MACT).

For both the NSPS and the emission guidelines, facilities are grouped into subcategories based on waste

burning capacity. Facilities whose capacities are less than or equal to 200 lb/hr are considered small facilities, those

whose capacity is greater than 200 lb/hr but less than or equal to 500 lb/hr are considered medium facilities, and those

whose capacity is greater than 500 lb/hr are considered large facilities. Separate emission limits apply to each

subcategory.

The NSPS establish standards that limit emissions from new MWIs. The standards are expected to reduce

mercury emissions by 45 to 75%. The NSPS also require training and qualification of MWI operators, incorporate siting

requirements, specify testing and monitoring requirements to demonstrate compliance with the emission limits, and

establish reporting and record keeping requirements.

The emission guidelines require States to develop regulations that limit emissions from existing MWIs. The

emission guidelines are expected to reduce emissions from existing MWIs by 93 to 95 percent. Consistent with the

NSPS, the emission guidelines also require training and qualification of MWI operators, specify testing and monitoring

requirements, and establish reporting and record keeping requirements. Existing MWIs would have to meet one of the

following two compliance schedules: (1) full compliance with an EPA-approved State plan within one year after approval

of the plan, or (2) full compliance with the State plan within three years after EPA approval of the State plan, provided

the State plan includes measurable and enforceable incremental steps of progress that will be taken to comply with the

have been used on some of the larger MWIs. These dry systems may use sorbent injection (e.g., lime)

via either dry injection or spray dryers upstream of the PM control device for acid gas control. All of

these systems have limited success in controlling mercury emissions. Recent U.S. EPA studies,

however, indicate that wet scrubbers as well as sorbent injection/fabric filtration systems can achieve

improved mercury control by adding activated carbon to the sorbent material (U.S. EPA, 1997a). (These

controls for MWIs are discussed in Volume VIII of this Report to Congress.)

The estimated mercury emission factors for MWIs were determined by Midwest Research

Institute from 172 emission tests on 59 facilities. An average emission factor was calculated using both

continuous and intermittent MWI’s. The average emission factor was weighted based on the distribution

of test runs for intermittent and continuous MWI’s, giving each test equal weight. Different control-type

dependent emission factors were also developed. All combustion controls and dry scrubbers without

carbon were assigned an emission factor of 3.70 x 10 , all wet scrubbers and fabric filters were assigned -5

an emission factor of 1.31 x 10 , and dry scrubbers with carbon were assigned an emission factor of 1.66 -6

x 10 (MRI, 1996). -6
4-30

Mercury emissions were estimated using incinerator capacity, control-type, and facility-type

information from EPA’s National Dioxin Emissions from Medical Waste Incinerators(U.S EPA,

1996). Emission estimates were calculated by first converting the charge rate (or incinerator capacity)

for each facility to waste burned per year. All facilities were assumed to operate at 2/3 of their capacity. 

Batch units were assumed to operate at 160 batches per year. Therefore, the charge rates of the batch

facilities (lb/batch) were multiplied by 2/3 and 160 to get waste burned per year. All commercial units

were assumed to operate at 2/3 of their capacity and 7776 hours per year. Therefore the charge rates of

the commercial facilities (lb/hr) were multiplied by 2/3 and 7776 to get pounds of waste burned per year. 

For all other non-batch, non-commercial facilities, the charge rate (lb/hr) was multiplied by 2/3 and the

hours/year for the facility type. The facility type, other than batch and commercial facilities, was

determined from the charge rate of the facility. Annual mercury emissions for each facility were

calculated by multiplying the waste burned per year by the appropriate emission factor for the facility’s

control type. The total 1996 annual mercury emissions were estimated to be 14.6 Mg (16.0 tons).

4.1.5 Hazardous Waste Combustors

For the purpose of this emissions inventory, hazardous waste combustors include hazardous

waste incinerators, lightweight aggregate kilns, and cement kilns permitted to burn hazardous waste. 

These hazardous waste burning cement kilns are not counted in the emissions estimate for Portland

Cement manufacturing in Section 4.2.2. 

Based on the U.S. EPA's 1995 emission estimates (U.S. EPA, 1995b), hazardous waste

combustors currently combine to emit a total of 6.4 Mg/year (7.1 tons/year) of mercury. Of this amount,

hazardous waste incinerators are estimated to emit 3.5 Mg/year (3.95 tons/year), or approximately 54

percent of the total, hazardous waste burning cement kilns are estimated to emit 2.7 Mg/year (2.9

tons/year), or about 42 percent, and lightweight aggregate kilns are estimated to emit 0.28 Mg/year (0.31

tons/year), or about 4 percent of the total.

4.1.5.1 Hazardous Waste Incinerators

A hazardous waste incinerator is an enclosed, controlled flame combustion device that is used to

treat primarily organic and/or aqueous waste, although some incinerators burn spent or unusable

ammunition and/or chemical agents. These devices may be fixed (in situ) or mobile (such as those used

for site remediation). Major incinerator designs include rotary kilns, liquid injection incinerators,

fluidized bed incinerators and fixed hearth incinerators.

Currently, 162 permitted or interim status incinerator facilities, having 190 units, are in operation

in the U.S. According to the U.S. EPA's List of Hazardous Waste Incinerators (November 1994),

another 26 facilities are proposed (i.e., new facilities under construction or in the process of being

permitted). Of the 162 facilities, 21 are commercial sites that burn about 700,000 tons of hazardous

waste annually. The remaining 141 are onsite or captive facilities that burn about 800,000 tons of waste

annually.

Hazardous waste incinerators are equipped with a wide variety of air pollution control devices. 

Typical devices include packed towers, spray dryers, or dry scrubbers for acid gas (e.g., HCl, Cl )2

control, as well as venturi scrubbers, wet or dry ESPs or fabric filters for particulate control. Most

incinerators use wet systems to scrub acid emissions (three facilities use dry scrubbers). Activated

carbon injection for controlling dioxin and mercury is being used at only one incinerator. New control

technologies, such as catalytic oxidizers and dioxin/furan inhibitors, have recently emerged but have not
4-31

Major Designs for Hazardous Waste Incinerators

Rotary Kilns. Rotary kiln systems typically contain two incineration chambers: the rotary kiln and an

afterburner. The shell of the kiln is supported by steel trundles that ride on rollers, allowing the kiln to rotate around

its horizontal axis at a rate of one to two revolutions per minute. Wastes are fed directly at one end of the kiln and

heated by primary fuels. Waste continues to heat and burn as it travels down the inclined kiln, which typically

operates at 50-200 percent excess air and at temperatures of 1600-1800 F. Flue gas from the kiln is routed to an o

afterburner, operating at 100-200 percent excess air and 2000-2500 F, where unburnt components of the kiln flue gas o

are more completely combusted. Some rotary kiln incinerators, known as slagging kilns, operate at high enough

temperatures that residual materials leave the kiln in molten slag form. The molten residue is then water-quenched. 

Ashing kilns operate at a lower temperature, with the ash leaving as a dry material.

Liquid Injection Incinerators. A liquid injection incineration system consists of an incineration chamber,

waste burner and auxiliary fuel system. Liquid wastes are atomized as they are fed into the combustion chamber

through waste burner nozzles.

Fluidized Bed Incinerators. A fluidized bed system is essentially a vertical cylinder containing a bed of

granular material at the bottom. Combustion air is introduced at the bottom of the cylinder and flows up through the

bed material, suspending the granular particles. Waste and auxiliary fuels are injected into the bed, where they mix

with combustion air and burn at temperatures from 840-1500 F. Further reaction occurs in the volume above the bed o

at temperatures up to 1800 F. o

Fixed Hearth Incinerators. These systems typically contain a primary and a secondary furnace chamber. 

The primary chamber operates in "starved air" mode and the temperatures are around 1000 F. The unburnt o

hydrocarbons reach the secondary chamber where 140-200 percent excess air is supplied and temperatures of 1400-

2000 F are achieved for more complete combustion. o

been used on any full-scale incinerators in the U.S.

4.1.5.2 Lightweight Aggregate Kilns

The term lightweight aggregate refers to a wide variety of raw materials (such as clay, shale or

slate) that after thermal processing can be combined with cement to form concrete products. Lightweight

aggregate concrete is produced either for structural purposes or for thermal insulation purposes. A

lightweight aggregate plant is typically composed of a quarry, a raw material preparation area, a kiln, a

cooler and a product storage area. The material is taken from the quarry to the raw material preparation

area and from there is fed into the rotary kiln.

There are approximately 36 lightweight aggregate kiln locations in the U.S. Of these sites, there are

currently seven facilities that burn hazardous waste in a total of 15 kilns.

Lightweight aggregate kilns use one or a combination of air pollution control devices, including

fabric filters, venturi scrubbers, spray dryers, cyclones and wet scrubbers. All of the facilities utilize

fabric filters as the main type of emissions control, although one facility uses a spray dryer, venturi

scrubber and wet scrubber in addition to a fabric filter.
4-32

Major Design and Operating Features of

Lightweight Aggregate Kilns

Rotary kilns at lightweight aggregate plants typically consist of a long (30 to 60-meter) steel cylinder lined

with refractory bricks. The cylinder is capable of rotating about its axis and is inclined at an angle of about 5 degrees.

Prepared raw material is fed into the kiln at the higher end, while firing takes place at the lower end. The

dry raw material fed into the kiln is initially preheated by hot combustion gases. Once the material is preheated, it

passes into a second furnace zone where it melts to a semiplastic state and begins to generate gases that serve as a

bloating or expanding agent. In this zone, specific compounds begin to decompose and form gases (such as SO , CO , 2 2

SO , and O ) that eventually trigger the desired bloating action within the material. As temperatures reach their 3 2

maximum (approximately 2100 F), the semiplastic raw material becomes viscous and entraps the expanding gases. o

This bloating action produces small, unconnected gas cells, which remain in the material after it cools and solidifies. 

The product exits the kiln and enters a section of the process where it is cooled with cold air and then conveyed to the

discharge.

4.1.5.3 Hazardous Waste Burning Cement Kilns

The process of burning hazardous waste in cement kilns differs from the combustion of nonhazardous waste only in the type of fuel used. For a complete discussion of the process, refer to Section

4.2.2. 

Emissions from cement kilns permitted to burn hazardous waste were derived by EPA for the 41

hazardous waste burning cement kilns in the United States. The data used to make the estimates was

supplied from the EPA Office of Solid Waste for the proposed hazardous waste combustion MACT

standards (U.S. EPA, 1997a). The national annual mercury estimate is 2.66 Mg/year (2.93 tons/year).

4.1.6 Residential Boilers

Residential boilers are relatively small boilers used in homes and apartments. These boilers may

use coal, oil, or natural gas as fuels; however, mercury emissions from natural gas combustion are

negligible. As with the other types of boilers, mercury vaporizes during combustion in the coal- and oilfired residential boilers and the emissions appear as a trace contaminant in the exhaust gas.

The estimated annual mercury emissions from residential boilers, 3.3 Mg/yr (3.6 tons/yr), are

related to the amount of fuel used in the combustion process. Estimates of coal and oil consumption

from these boilers on a per-state basis are presented in Table A-5, Appendix A. Residential boilers

consume energy at an annual rate of 6.2 x 10 MJ/yr (5.9 x 10 Btu/yr). About 1 percent of this energy 12 15

consumption results from coal combustion, 15 percent from oil and petroleum fuel combustion and

85 percent from natural gas combustion (U.S. Department of Energy, 1996).

Because there is no evidence to link mercury emissions to boiler type, this section does not

describe residential boiler types. Information on boiler types may be found in the Air Pollution

Engineering Manual, AP-42 and the L&E document (Buonicore and Davis, 1992; U.S. EPA, 1988; U.S.

EPA, 1997a).

Estimated mercury emission factors for coal combustion in residential boilers are the same as
4-33

those used for other coal combustion processes. These calculations include the assumption that all

mercury fired with the coal is emitted as stack gas. An estimated emission factor of 7.0 kg/10 J 15

(16 lb/10 Btu) was used for bituminous coal combustion, and 7.6 kg/10 J (18 lb/10 Btu) was used for 12 15 12

anthracite coal combustion. Estimates of mercury emissions on a per-state basis from coal-fired

residential boilers were determined by using these emission factors and the coal consumption estimates

for the states as presented in Table A-5, Appendix A. These calculated emission values are presented in

Table A-6, Appendix A. In estimating emissions, it was assumed that mercury emissions from

residential boilers were not controlled. The total annual estimated emissions for coal-fired residential

boilers is 0.4 Mg/yr (0.5 tons/yr).

The estimated mercury emissions for oil combustion were estimated by using an emission factor

of 2.9 kg/10 J (6.8 lb/10 Btu) for residual oil and 3.0 kg/10 J (7.2 lb/10 Btu) for distillate oil and 15 12 15 12

the oil consumption estimates for the states given in Table A-5, Appendix A. These estimated emissions

values are presented in Table A-7, Appendix A. The total annual estimated emissions for oil-fired

residential boilers is 2.9 Mg/yr (3.2 tons/yr).

4.1.7 Sewage Sludge Incinerators

Sewage sludge incinerators (SSIs) are operated primarily by U.S. cities and towns as a final stage

of the municipal sewage treatment process. The locations of SSIs in the United States are given in Figure

4-10. The mercury in sewage comes from households, commercial and industrial sources and industries

discharging industrial wastewater into the sewer systems and flows to sewage treatment plants. After

treatment at the sewage treatment plant, the sludge is usually land filled or incinerated. Only a small

percentage of U.S. cities use sewage sludge incinerators. The estimated annual mercury emissions in

1994 from SSIs account for 0.86 Mg/yr (0.94 tons/yr). Mercury emissions occur when mercury in the

sewage is combusted at high temperatures, vaporizes and exits through the gas exhaust stack. Land filled

sludge or sludge applied to farmland are also potential sources of mercury emissions. These sources are

not addressed in this inventory.

A total of 116 SSIs currently operate in the United States. An estimated 785,000 Mg

(865,000 tons) of sewage sludge on a dry basis are incinerated annually (U.S. EPA, 1993b). Most

facilities are located in the Eastern United States, but a substantial number also are located on the West

Coast. New York has the largest number of SSI facilities with 33, followed by Pennsylvania and

Michigan with 21 and 19, respectively.

Within the SSI category, three combustion techniques are used: multiple-hearth, fluidized-bed

and electric infrared. Multiple-hearth units predominate; over 80 percent of the identified SSIs are

multiple hearth. About 15 percent of the SSIs in operation are fluidized bed units, about 3 percent are

electric infrared and the remainder co-fire sewage sludge with municipal waste (U.S. EPA, 1993b).

The sewage sludge incinerator process involves two primary steps: dewatering the sludge and

incineration. The primary source of mercury emissions from SSIs is the combustion stack. Most SSIs

are equipped with some type of wet scrubbing system for PM control. Because wet systems provide gas

cooling, as well as PM removal, these systems can potentially provide some mercury control.
Figure 4-10

Sewage Sludge Incinerators

IThe U.S. EPA’s Compilation of Air Pollutant Emission Factors (U.S. EPA, 1988a) (otherwise

known as the AP-42) for SSIs lists five mercury emission factors for various types of SSIs and controls:

0.005 g/Mg (1.0 x 10” lb/ton) for multiple hearth combustors controlled with a combination of venturi

and impingement scrubbers, 0.03 g/Mg (6.0 x 10e5 lb/ton) for fluidized bed combustors controlled with

a combination of venturi and impingement scrubbers, 2.3 g/Mg (4.6 x 10” lb/ton) for multiple hearth

combustors controlled with a cyclone scrubber, 1.6 g/Mg (3.2 x 10” lb/ton) for multiple hearth

combustors controlled with a combination of cyclone and venturi scrubbers, and 0.97 g/Mg (1.94 x lo3 lb/ton) for multiple hearth combustors controlled with an impingement scrubber (U.S. EPA, 1993b).

Given that combustor and control types are not known for all SSIs currently operating in the United

States, average emission factors were calculated: 0.0175 g/Mg (3.5 x 10” lb/ton) for SSIs controlled

with a combination of venturi and impingement scrubbers and 1.623 g/Mg (3.25 x 10e3 lb/ton) for SSIs

controlled by any other type or combination of types of scrubbers. Of the SSIs where data are

available, 32.6 percent of SSIs are controlled by a combination of venturi and impingement scrubbers,

and 67.4 percent are controlled by some other means. These percentages were assumed to apply to the

total population of SSIs.. Multiplying the total amount of sewage sludge incinerated annually,

785,000 Mg (865,000 x lo6 tons), by the appropriate percentage and emission factor gives a mercury

emission estimate of 4.5 x 10” Mg/yr (4.9 x 10” tons/yr) for SSIs controlled with a combination of

venturi and impingement scrubbers and an estimate of 0.86 Mg/yr (0.94 tons/yr) for SSIs controlled by

some other means. The overall mercury emissions estimate from SSIs is, thus, 0.86 Mg/yr

(0.94 tons/yr).

4-34
4-35

4.1.8 Wood Combustion

Wood and wood wastes are used as fuel in both the industrial and residential sectors. In the

industrial sector, wood waste is fired in industrial boilers to provide process heat, while wood is used in

fireplaces and wood stoves in the residential sectors. Studies have shown that wood and wood wastes

may contain mercury. Insufficient data are available, however, to estimate the typical mercury content

of wood and wood wastes.

Wood waste combustion in boilers is mostly confined to industries in which wood waste is

available as a byproduct. These boilers, which are typically of spreader stoker or suspension-fired

design, are used to generate energy and alleviate possible solid waste disposal problems. In boilers,

wood waste is normally burned in the form of hogged wood, sawdust, shavings, chips, sanderdust, or

wood trim. Heating values for this waste range from about 9,300 to 12,000 kJ/kg (4,000 Btu/lb to

5,000 Btu/lb) of fuel on a wet, as-fired basis. The moisture content is typically near 50 weight percent

but may vary from 5 to 75 weight percent, depending on the waste type and storage operations. As of

1980, about 1,600 wood-fired boilers were operating in the United States, with a total capacity of

approximately 30.5 gigawatts (GW) (1.04 x 10 Btu/hr) (U.S. EPA, 1982). No specific data on the 11

distribution of these boilers were identified but most are likely to be located where pulp and paper plants

or logging operations are located (i.e., in the Southeast, the Pacific Northwest States, Wisconsin,

Michigan, and Maine) (U.S. EPA, 1993a). One National Council of the Paper Industry for Air and

Stream Improvement (NCASI) study found the mercury content of bark waste to range from <0.08 to

0.84 ppm by weight (NCASI, 1991).

Wood-fired boilers use PM control equipment, which may provide some reduction in mercury

emissions. The most common control devices used to reduce PM emissions from wood-fired boilers are

mechanical collectors, wet scrubbers, ESPs, and fabric filters. Only the last three have the potential for

mercury reduction. The most widely used wet scrubbers for wood-fired boilers are venturi scrubbers,

although no data have been located on the performance of these systems relative to mercury emissions. 

No data are available on mercury emission reduction for fabric filters for wood combustors, but results

for other combustion sources suggest that efficiencies will be low, probably 50 percent or less (U.S.

EPA, 1997a).

The data on mercury emissions from wood-fired boilers are limited. A recent AP-42 study

provided a range and average typical emission factor for wood waste combustion in boilers based on the

results of seven tests. The average emission factor of 2.6 x 10 kg/Mg (5.2 x 10 lb/ton) of wood burned -6 -6

is recommended as the best typical emission factor for wood waste combustion in boilers (U.S. EPA,

1992c). Dividing the total capacity of wood-fired boilers, 30.5 GW (1.04 x 10 Btu/hr), by the average 11

heating value of wood, 10,600 kJ/kg (4,560 Btu/lb), gives the total hourly rate: 10,367 Mg/hr

(11,404 tons/hr) (U.S. EPA, 1996). Assuming that wood-fired boilers operate at capacity at 8,760 hr/yr

and multiplying by the above emission factor gives a mercury emission estimate for wood-fired boilers

of 0.24 Mg/yr (0.26 tons/yr). This estimate has a high degree of uncertainty given the limited data

available.

Wood stoves, which are commonly used as residential space heaters, are of three different types: 

(1) the conventional wood stove, (2) the noncatalytic wood stove and (3) the catalytic wood stove. 

Fireplaces are used primarily for aesthetic effects and secondarily as a supplemental heating source in

homes and other dwellings. Wood is most commonly used as fuel, but coal and densified wood "logs"

also may be burned.
4-36

All of the systems described above operate at temperatures that are above the boiling point of

mercury. Although some wood stoves use emission control measures to reduce volatile organic

compound (VOC) and carbon monoxide (CO) emissions, these techniques are not expected to affect

mercury emissions. Consequently, any mercury contained in the wood will be emitted with the

combustion gases via the exhaust stack.

For residential wood combustion, only one emission factor, 1.3 x 10 kg/Mg (2.6 x 10 lb/ton) is -2 -2

available, which is based on a single test burning a single type of wood (pine) at a single location

(DeAngelis et al., 1980). In 1987, the Department of Energy estimated that 22.5 million households

burned approximately 42.6 million cords of wood (Phillips, 1993). Given that the densities of wood vary

greatly depending on wood type and the moisture content of the wood, and because the above emission

factor is from a single test, nationwide emissions of mercury for residential wood combustion were not

estimated.

4.1.9 Crematories

Volatilization of mercury from the mercury alloys contained in amalgam tooth fillings during

cremation of human bodies is a potential source of mercury air emissions. In 1995, there were 488,224

cremations in the 1,155 crematories located throughout the United States (Cremation Association of

North America, 1996). 

Only one set of data are available for the average quantity of mercury emitted for a cremation in

the United States. Tests were conducted for a propane-fired incinerator at a crematorium in California. 

Results of the testing for uncontrolled mercury emissions ranged from 3.84 x 10 to 1.46 x 10 kg/body -8 -6

burned (8.45 x 10 to 3.21 x 10 lb/body); the average mercury emission factor was 0.94 x 10 kg/body -8 -6 -6

burned (2.06 x 10 lb/body). The test results were obtained from a confidential test report to the -6

Califonia Air Resource Board (FIRE, 1995).

Multiplying the number of cremations in the United States by the average emission factor results

in 1995 annual mercury emissions of 4.6 x 10 Mg (5.1 x 10 tons). -4 -4

4.2 Manufacturing Sources

Manufacturing sources, including processes that use mercury directly and those that produce

mercury as a byproduct, account for an estimated 14.4 Mg/yr (15.6 tons/yr) of mercury emissions

generated in the United States. Emissions from these sources are presented in Table 4-9 and are

discussed below.

4.2.1 Chlor-alkali Production Using the Mercury Cell Process

Chlor-alkali production using the mercury cell process, which is the only chlor-alkali process

using mercury, accounted for 14.7 percent of all U.S. chlorine production in 1993 (Dungan, 1994). 

Although most chlor-alkali plants use diaphragm cells, the mercury cell is still in use at some facilities. 

Each mercury cell may contain as much as 3 tons of mercury, and there are close to 100 cells at each

mercury cell plant, making chlor-alkali plants a well-known source of mercury release. As new plants 

and/or additional capacity is added, however, the chlor-alkali industry is moving away from mercury cell

production and toward a membrane cell process because the membrane cell process does not use mercury

and is more energy efficient than the mercury cell process (Rauh, 1991). Companies have 
4-37

Table 4-9

Best Point Estimates of Mercury Emissions from Anthropogenic Manufacturing Sources: 1994-1

Source Mg/yr Tons/yr % of total Date of Data Uncertainty Basis for Em

Emissions

a

Degree of

b

Chlor-alkali production 6.5 7.1 4.5 1994/1994 Medium Section 114 industry sPortland cement manufacturing 4.4 4.8 3.1 /1994 Medium Test reports; Industry 

category are 3.3 tons/y

Pulp and paper manufacturing 1.7 1.9 1.2 /1994 High Test dataInstrument manufacturing 0.5 0.5 0.3 1973/1992 High Survey questionnaire rSecondary mercury production 0.4 0.4 0.3 1997/1994 High TRI data Electrical apparatus manufacturing 0.3 0.3 0.2 1973/1996 High Engineering judgmentCarbon black production 0.3 0.3 0.2 1980/1995 High Test dataLime manufacturing 0.1 0.1 0.1 1986/1994 High Test data and mass baPrimary lead smelting 0.1 0.1 0.1 1993/1994 High Test dataPrimary copper smelting 0.06 0.06 0.0 1994/1994 High Test reports and enginFluorescent lamp recycling 0.005 0.006 0.0 1993/1993 High Test data and mass baBattery production 0.0005 0.0006 0.0 1986/1995 High Engineering judgmentPrimary mercury production - - - - - Insufficient data to estMercury compounds production - - - - - Insufficient data to estByproduct coke production - - - - - Insufficient data to estPetroleum refining - - - - - Insufficient data to estTotal 14.4 15.6 10.0

 Date that data emission factor is based on/Date of activity factor used to estimate emissions. a

 A "medium" degree of uncertainty means the emission estimate is believed to be accurate within + 25 percent. A "high" degree of uncertainty means 

b

believed to be accurate within + 50 percent.
4-38

been waiting until major capital investments are required for current installations before converting to

processes that do not use mercury. When chlor-alkali plants replace mercury cells with alternative

technologies, thousands of tons of mercury have to be disposed of as hazardous waste. There is currently

no approved disposal method for mercury; only recovery/recycling of mercury is currently allowed under

RCRA.

Table 4-10 lists U.S. mercury-cell chlor-alkali production facilities and their capacities. Figure

4-11 shows the location of these facilities across the U.S. The chlor-alkali industry is the largest user of

mercury; however, the amount of chlorine produced using mercury cells has declined over the past 20

years (Cole et al., 1992). According to the Chlorine Institute, there are 14 chlor-alkali plants that

currently use mercury cells compared to 25 facilities, 20 years ago (The Chlorine Institute, 1991). There

are no plans for construction of new mercury-cell chlor-alkali facilities (Rauh, 1991).

The three primary sources of mercury air emissions are the (1) byproduct hydrogen stream,

(2) end box ventilation air and (3) cell room ventilation air. The byproduct hydrogen stream from the

decomposer is saturated with mercury vapor and may also contain fine droplets of liquid mercury. The

quantity of mercury emitted in the end box ventilation air depends on the degree of mercury saturation

and the volumetric flow rate of the air. The amount of mercury in the cell room ventilation air is variable

and comes from many sources, including end box sampling, removal of mercury butter from end boxes,

maintenance operations, mercury spills, equipment leaks, cell failure, and other unusual circumstances

(U.S. EPA, 1984).

Mercury cell chlor-alkali facilities use pollution prevention methods to minimize mercury

emissions to the environment. In the United States many facilities are installing thermal desorption or

alternate technology to reduce mercury discharges to land (hazardous waste disposal sites). The amount

of training provided to employees and the number of inspections have been increased to reduce the

possibilities of mercury releases. In addition, equipment has been upgraded to reduce the likelihood of

mercury spills (The Chlorine Institute, 1991).

The control techniques that are typically used to reduce the level of mercury in the hydrogen

streams and in the ventilation stream from the end boxes are these: (1) gas stream cooling, (2) mist

eliminators, (3) scrubbers, and (4) adsorption on activated carbon or molecular sieves. Mercury

emissions via the cell room air circulation are not subject to specific emission control measures. 

Concentrations are maintained, however, at acceptable worker exposure levels through good

housekeeping practices and equipment maintenance procedures (U.S. EPA, 1984).
4-39

Table 4-10

1996 U.S. Mercury-Cell Chlor-Alkali Production Facilitiesa

Facility Location

Capacity,

10 Mg/yr 3

Capacity,

10 tons/yr 3

1994

emissionsb

Mg/yr

Georgia-Pacific Corp., Chemical

Division

Bellingham, WA 82 90 0.585

BF Goodrich, Chemical Group Calvert City, KY 109 120 0.382

Hanlin Group, Inc., LCP

Chemicals Division

Reigelwood, NC

Orrington, ME

48

76

53

80

0.497

0.264

ASHTA Chemicals, Inc. Ashtabula, OH 36 40 0.753

Occidental Petroleum

Corporation, Electrochemicals

Division

Deer Park, TX

Delaware City, DE

Muscle Shoals, AL

347

126

132

383

139

146

0.472

0.231

0.106

Olin Corporation, Olin

Chemicals

Augusta, GA

Charleston, TN

102

230

112

254

0.597

0.684

Pioneer Chlor Alkali Company,

Inc.

St. Gabriel, LA 160 176 N/Ac

PPG Industries, Inc., Chemicals

Group

Lake Charles, LA

New Martinsville, WV

233

70

256

77

0.558

0.513

Vulcan Materials Company,

Vulcan Chemicals Division

Port Edwards, WI 65 72 N/Ac

TOTAL 1,816 1,998 6.48 c

(7.14

tons/yr)

 SRI International, 1996 a

 TRI emissions data (EPA, 1996b). b

N/A = Not available from survey questionnaires. For the purposes of this inventory, it is assumed that facilities not c 

reporting mercury emissions emitted the average of the other facilities. These assumed values are reflected in

the total.
Figure 4-11

Chlor-Alkali Production Facilities

The Mercury-Cell Chlor-Alkali Process

The mercury-cell chlor-alkali process consists of two electrochemical ceils, the electrolyzer and the

decomposer. A purified solution of saturated sodium or potassium brine flows from the main brine saturation section,

through the inlet end box and into the electrolyzer. The brine flows between stationary activated titanium anodes

suspended in the brine from above and a mercury cathode, which flows concurrently with the brine over a steel base

(U.S. EPA, 1984).

Chlorine gas is formed at the electrolyzer anode and is collected for further treatment. The spent brine is

recycled from the electrolyzer to the main brine saturation section through a dechlorination stage. Sodium is collected

at the electrolyzer cathode, forming an amalgam containing from 0.25 to 0.5 percent sodium. The outlet end box

receives the sodium amalgam from the electrolyzer, keeping it covered with an aqueous layer to reduce mercury

emissions. The outlet end box also allows removal of thick mercury “butter” that is formed through the outlet end box

into the second cell (the decomposer) (U.S. EPA, 1984).

The decomposer is a short-circuited electrical cell in an electrolytic sodium hydroxide solution. This cell has

the sodium amalgam as the anode and graphite or metal as the cathode. Water added to the decomposer reacts with the

sodium amalgam to produce elemental mercury, sodium hydroxide and hydrogen gas (a byproduct). The mercury,

stripped of sodium, is recirculated to the cell through the inlet end box. The caustic soda solution typically leaves the

decomposer at a concentration of 50 percent (by weight) and is filtered and further concentrated by evaporation. The

byproduct hydrogen gas may be vented to the atmosphere, burned as a fuel, or used as a feed material for other

processes (U.S. EPA, 1984).

4-40
4-41

Gas stream cooling may be used as the primary mercury control technique or as a preliminary

removal step to be followed by a more efficient control device. The hydrogen gas stream from the

decomposer exits at 93 to 127C (200 to 260F) and passes into a primary cooler. In this indirect cooler,

a shell-and-tube heat exchanger with ambient temperature water is used to cool the gas stream to 32 to

43C (90 to 110F). A knockout container following the cooler is used to collect the mercury. If

additional mercury removal is desired, the gas stream may be passed through a more efficient cooler or

another device. Direct or indirect coolers using chilled water or brine provide for more efficient mercury

removal by decreasing the temperature of the gas stream to 3 to 13C (37 to 55F). Regardless of the

gas stream treated, the water or brine from direct contact coolers requires water treatment prior to reuse

or discharge because of the dissolved mercury in the liquid (U.S. EPA, 1984).

Mist eliminators (most commonly the filter pad type) can be used to remove mercury droplets,

water droplets, or PM from the cooled gas streams. Particles trapped by the pad are removed by

periodically spraying the pad and collecting and treating the spray solution (U.S. EPA, 1984).

Scrubbers are used to absorb the mercury chemically from both the hydrogen stream and the end

box ventilation streams. The scrubbing solution is either depleted brine from the mercury cell or a

sodium hypochlorite (NaOCl) solution. These solutions are used in either sieve plate scrubbing towers

or packed-bed scrubbers. Mercury vapor and mist react with the sodium chloride or hypochlorite

scrubbing solution to form water-soluble mercury complexes. If depleted brine is used, the brine

solution is transferred from the scrubber to the mercury cell, where it is mixed with fresh brine, and the

mercury is recovered by electrolysis in the cell (U.S. EPA, 1984).

Sulfur- and iodine-impregnated carbon adsorption systems are commonly used to reduce the

mercury levels in the hydrogen gas stream if high removal efficiencies are desired. This method requires

pretreatment of the gas stream by primary or secondary cooling followed by mist eliminators to remove

about 90 percent of the mercury content of the gas stream. As the gas stream passes through the carbon

adsorber, the mercury vapor is initially adsorbed by the carbon and then reacts with the sulfur or iodine

to form the corresponding mercury sulfides or iodides. Several adsorber beds in series can be used to

reduce the mercury levels to the very low parts per billion (ppb) range (U.S. EPA, 1984).

Mercury emissions data from chlor-alkali facilities were obtained from Clean Air Act

section 114 survey questionnaires (BF Goodrich, 1992; Georgia-Pacific, 1993; LCP Chemicals, 1993a;

LCP Chemicals, 1993b; Occidental, 1993; Olin Chemicals, 1993a; Olin Chemicals, 1993b; Pioneer Chlor

Alkali, 1993; PPG Industries, 1993a; PPG Industries, 1993b; Vulcan Materials, 1993). The data reported

are for 1991. Data are also available from the Toxic Release Inventory (TRI) (U.S. EPA, 1996). The

estimated mercury emissions were 6.5 Mg (7.1 tons) and included reported mercury emissions from 12

of the 14 mercury cell chlor-alkali production facilities listed in Table 4-11. For the purposes of this

inventory, the two remaining facilities (Vulcan Materials and Pioneer Chlor Alkali) were assumed to

emit the average of the other 12 facilities because reported data were not available from either the CAA

section 114 survey questionnaires or the 1996 TRI. 

 4.2.2 Cement Manufacturing

United States cement kiln capacity data for 1990 showed a total of 212 U.S. cement kilns with a

combined total capacity of 73.5 x 10 Mg (81 x 10 tons) (U.S. EPA, 1993a). Of this total, 201 kilns 6 6

were active and had a total clinker capacity of 71.8 x 10 Mg (79.1 x 10 tons) (U.S. EPA, 1993a). 6 6

Because the majority (96 percent) of this cement was portland cement, portland cement production
processes and emissions are the focus of this section (U.S. EPA, 1993a). Total mercury emissions

from the portland cement process are estimated to be 4.4 Mg (4.8 tons) per year. In 1990, 68 percent

of portland cement was produced by the dry process and 32 percent by the wet process (Portland

Cement Association, 1991). The locations of active cement manufacturing plants in the continental

U.S. are shown in Figure 4-12.

The primary sources of mercury emissions from Portland cement manufacturing are expected to

be from the kiln and preheating/precalcining steps. Small quantities of mercury may be emitted as a

contaminant in the PM from process fugitive emission sources. Process fugitive emission sources

include materials handling and transfer, raw milling and drying operations in dry process facilities and

finish milling operations. Typically, PM emissions from these process fugitive sources are captured by

a ventilation system controlled with a fabric filter. No data are available on the ability of these systems

to capture mercury emissions from cement kilns.

In the pyroprocessing units, PM emissions are controlled by fabric filters and ESPs.. Clinker

cooler systems are controlled most frequently with pulse jet or pulse plenum fabric filters. No data are

available on the ability of these control systems to capture mercury emissions from cement kilns.

Figure 4-12

Cement Manufacturing Plants

4-42
4-43

Mercury present in the raw material and the fuel is likely to be emitted from all four cement

processes summarized in the text box. Cement kiln test reports were reviewed from a number of

facilities performing Certification of Compliance (COC) tests which are required of all kilns burning

waste-derived fuel (WDF). Emission tests from two other kilns were also reviewed in this analysis. In

all, 15 test runs provided enough information to calculate an emission factor (some of these were from

the same kiln). This information included clinker production as well as mercury emission rates and

process conditions.

The mercury emissions discussed in this section for the manufacture of portland cement are only

for the use of fossil fuels and nonhazardous waste auxiliary fuels; mercury emissions from the use of

hazardous waste fuels burned at cement manufacturing facilities are accounted for in the calculation of

mercury emissions from hazardous waste combustors (Section 4.1.5). 

The principal sources of mercury emissions are expected to be from the kiln and

preheating/precalcining steps. Negligible quantities of emissions would be expected in the raw material

processing and mixing steps because the only source of mercury would be fugitive dust containing

naturally occurring quantities of mercury compounds from the raw materials. Processing steps that occur

after the calcining process in the kiln would be expected to be a much smaller source of emissions than

the kiln. Potential mercury emission sources are denoted by solid circles in Figure 4-10. Emissions

resulting from all processing steps include particulate matter. Additionally, emissions from the

pyroprocessing step include other products of fuel combustion such as SO , NO , and CO. Carbon 2 x

dioxide from the calcination of limestone will also be present in the flue gas.

Cement kiln test reports have been reviewed by EPA (and its contractor) in its development of

the portland cement industry NESHAP, and by a private company. Test reports for Certification of

Compliance (COC) emissions tests (required of all kilns burning hazardous waste derived fuel) and test

reports for facilities not burning hazardous waste (RTI, 1996; Gossman, 1996) were reviewed. The

results from the Gossman study showed and average emission factor of 0.65 x 10 kg/Mg of clinker (1.3 -4

x 10 lb/ton of clinker) for nonhazardous waste fuels The RTI study evaluated tests based on both -4

nonhazardous waste fuel and hazardous waste fuel. For the hazardous waste tests, the mercury emissions

data were corrected to reflect only the mercury emissions originating from the fossil fuel and raw

material. The emissions data for nonhazardous waste and the corrected hazardous waste were combined

and showed an average mercury emission factor of 0.65 x 10 kg/Mg of clinker (1.29 x 10 lb/ton of -4 -4

clinker).

4.2.3 Pulp and Paper Manufacturing

In the pulp and paper industry, wood pulp is produced from raw wood via chemical or

mechanical means or a combination of both. When chemical pulping methods are used to produce pulp,

the chemicals used in the process are recycled for reuse in the process. Combustion sources located in

the chemical recovery area of pulp and paper mills represent potential sources of mercury emissions. 

Four principal chemical wood pulping processes currently in use are (1) kraft, (2) soda, (3)

sulfite, and (4) semichemical. (The semichemical process requires both chemical and mechanical

treatment of the wood.) The kraft process is the dominant pulping process in the United States,

accounting for approximately 80 percent of the domestic pulp production. Currently, there are estimated

to be 122 kraft, 2 soda, 15 sulfite, and 14 stand-alone semichemical pulp mills in the United States with

chemical recovery combustion (Nicholson, 1996; Soltis, 1995; McManus, 1996).
4-44

The Portland Cement Manufacturing Process

The portland cement manufacturing process can be divided into four major steps: raw material

acquisition and handling, kiln feed preparation, pyroprocessing, and finished cement grinding (U.S. EPA,

1993a).

The initial step in the production of portland cement manufacturing is acquiring raw materials,

including limestone (calcium carbonate) and other minerals such as silica.

Raw material preparation, the second step in the process, includes a variety of blending and sizing

operations designed to provide a feed with appropriate chemical and physical properties. Raw material

processing differs somewhat for the "wet" and "dry" processes. At dry process facilities, the moisture content

in the raw material, which can range between 2 and 35 percent, is reduced to less than 1 percent. Heat for

drying is often provided by the exhaust gases from the pyroprocessor (i.e., kiln). At facilities where the wet

process is used, water is added to the raw material during the grinding step, thereby producing a pumpable

slurry containing approximately 65 percent solids. 

Pyroprocessing (thermal treatment) of the raw material is carried out in a rotary kiln, which is the

heart of the Portland cement manufacturing process. During pyroprocessing, the raw material is transformed

into clinkers, which are gray, glass-hard, spherically shaped nodules that range from 0.32 to 5.1 cm (0.125 to

2.0 in.) in diameter.

The rotary kiln is a long, cylindrical, slightly inclined, refractory-lined furnace. The raw material

mix is introduced in the kiln at the elevated end, and the combustion fuels are introduced into the kiln at the

lower end, in a countercurrent manner. The rotary motion of the kiln transports the raw material from the

elevated end to the lower end. Fuel such as coal or natural gas (or occasionally oil) is used to provide energy

for calcination and sintering. Other fuels, such as shredded municipal garbage, chipped rubber, petroleum

coke, and waste solvents are also being used more frequently. Mercury is present in coal and oil and may

also be present in appreciable quantities in the waste-derived fuels mentioned above. Because mercury

evaporates at approximately 350C (660F), most of the mercury present in the raw materials may be emitted

during the pyroprocessing step. Combustion of fuel during the pyroprocessing step also contributes to

mercury emissions. Pyroprocessing can be accomplished by one of four different processes: wet process,

dry process, dry process with a preheater, and dry process with a preheater/precalciner. These processes

accomplish the same physical and chemical steps described above.

The last step in the pyroprocessing is cooling the clinker. This process step recoups up to 30 percent

of the heat input to the kiln system, locks in desirable product qualities by freezing mineralogy, and makes it

possible to handle the cooled clinker with conventional conveying equipment. Finally, after the cement

clinker is cooled, a sequence of blending and grinding operations is carried out to transform the clinker into

Due to state and federal regulations for PM emissions, almost all chemical recovery combustion

units at kraft pulp mills (i.e., recovery furnaces, smelt dissolving tanks, and lime kilns) are equipped with

add-on PM control devices. There are only limited emission test data from pulp and paper combustion

sources on the performance of these add-on controls for metals such as mercury. However, data

collected from other combustion sources on the relative performance of add-on control devices for

metals indicate that systems that achieve the greatest PM removal also provide the best performance for

metals. Therefore, particulate mercury may also be controlled to the same extent as PM. Although no

data are available for confirmation, some of the mercury may be emitted from the control devices in

vapor form, especially from the electrostatic precipitators, which have higher outlet temperatures

compared to wet scrubbers.
4-45

Mercury can be introduced into the pulping process through wood that is being pulped, in the

process water used in the pulping process, and as a contaminant in makeup chemicals added to the

process. If the mercury is not purged from the process in wastewater or as dregs, it can accumulate in the

chemical recovery area and subsequently be emitted from the chemical recovery combustion sources. 

The amount of mercury emitted may depend on the degree to which the pulping process is tightly closed

(i.e., the degree to which process waters are recycled and reused). 

Nearly all of the mercury emissions from pulp and paper manufacturing are from kraft and soda

recovery processes (approximately 99.9 percent) (U.S. EPA, 1997). To estimate the emissions, the firing

rate for each facility was multiplied by the emission factor for recovery furnaces (1.95x10 kg/Mg) -5

(Holloway, 1996). Estimated emissions from all of the facilities were then summed together to arrive at

the 1996 estimated mercury emissions of 1.7 Mg (1.9 tons) per year for the inventory as a whole.

4.2.4 Instrument (Thermometers) Manufacturing

Mercury is used in many medical and industrial instruments for measurement and control

functions. These instruments include thermometers, pressure-sensing devices and navigational devices. 

In 1992, an estimated 0.5 Mg (0.5 ton) of mercury was emitted from instrument manufacture; however,

this estimate should be used with caution as discussed below.

It is beyond the scope of this study to discuss all instruments that use mercury in some

measuring or controlling function. Although there is potential for mercury emissions from all

instruments containing mercury, this section focuses only on the production of thermometers because

they represent the most significant use, are usually disposed of in household waste (U.S. EPA, 1992a),

and more information is available on thermometer manufacture than on the manufacture of other

instruments.

There are generally two types of clinical thermometers: 95 percent are oral/rectal/baby

thermometers, and 5 percent are basal (ambient air) temperature thermometers. An oral/rectal/baby

thermometer contains approximately 0.61 grams of mercury and a basal thermometer contains

approximately 2.25 grams (U.S. EPA, 1992a).

During the production of thermometers, mercury emissions can be generated from mercury

purification and transfer, the mercury filling process, the heating-out/burning-off steps, and accidents

including spills of mercury and broken thermometers (U.S. EPA, 1997a). Within the industry, vapor

emissions from mercury purification and transfer are typically controlled by containment procedures,

local exhaust ventilation, temperature reduction to reduce the vapor pressure, dilution ventilation, or

isolation of the operation from other work areas. The bore sizing step can be modified to reduce the use

of mercury and be performed in an isolated room. Other measures that may be applied to this step are

use of local exhaust ventilation, dilution ventilation and temperature control (U.S. EPA, 1997a).
4-46

The Glass Thermometer Manufacturing Process

The production of glass thermometers begins by cutting glass tubes into required lengths and bore

sizes. Next, either a glass or metal bulb, used to contain the mercury, is attached to the base of the tube. The

tubes are filled with mercury in an isolated room. A typical mercury filling process is conducted inside a bell

jar. Each batch of tubes is set with open ends down into a pan, and the pan set under the bell jar, which is

lowered and sealed. The tubes are heated to approximately 200C (390F), and a vacuum is drawn inside the

bell jar. Mercury is allowed to flow into the pan from either an enclosed mercury addition system or a

manually filled reservoir. When the vacuum in the jar is released, the resultant air pressure forces the

mercury into the bulbs and capillaries. After filling, the pan of tubes is manually removed from the bell jar. 

Excess mercury in the bottom of the pan is refiltered and used again in the process (Reisdorf and D'Orlando,

1984).

Excess mercury in the tube stems is forced out the open ends by heating the bulb ends of the tubes in

a hot water or oil bath. The mercury column is shortened to a specific height by flame-heating the open ends

(burning-off process). The tubes are cut to a finished length just above the mercury column, and the ends of

the tubes are sealed. All of these operations are performed manually at various work stations. A temperature

scale is etched onto the tube, completing the assembly (Reisdorf and D'Orlando, 1984).

Disposal of thermometers also may result in releases. There are currently no recycling efforts

underway for mercury thermometers. The long life and small number of thermometers make a recycling

effort impracticable. Mercury thermometers enter the waste stream by being discarded from residential

and clinical settings. The thermometer is usually cracked or broken. In 1989, an estimated 16.3 tons of

mercury were discarded in thermometers, or just over 2 percent of total discards of mercury (Kiser,

1991). No information was available on how much of that total was land filled as opposed to incinerated

or the emissions generated from each.

No specific data for mercury emissions from manufacturing thermometers or any other

instrument containing mercury were found in the literature. One 1973 U.S. EPA report, however, 

presents an emission factor of 9 kg of mercury emitted for each megagram of mercury used (18 lb/ton) in

overall instrument manufacture (Anderson, 1973). This emission factor should be used with caution,

however, as it was based on survey responses gathered in the 1960s and not on actual test data. 

Instrument production and the mercury control methods used in instrument production have probably

changed considerably since the time of the surveys.

In 1992, 52 Mg (57 tons) of mercury was used in all instrument production (Anderson, 1973). 

Multiplying the emission factor above by the 1992 usage gives a mercury emission estimate of 0.5 Mg

(0.5 ton) for instrument manufacture. Again, a large degree of uncertainty is associated with this

estimate because of the concerns about the reliability in the emission factor.

Trends in mercury emissions from thermometer use and production are relatively stable. Since

1984, digital thermometers have begun to replace clinical mercury thermometers in clinics, hospitals and

doctors' offices. It is expected that this trend will continue. Mercury thermometers will continue to be

used in residential settings because of infrequent use and the higher cost for digital thermometers. The

decrease in mercury thermometer use attributable to the switch to digital thermometers in professional

settings will likely be offset by an increase in mercury thermometers purchased due to increased

population. The mercury content of thermometers will probably remain the same. Overall mercury

entering the waste stream from thermometers will likely remain stable (U.S. EPA, 1992a).
4-47

Secondary Mercury Production Processes

Secondary mercury production (recycling) can be accomplished by one of two general methods: chemical

treatment or thermal treatment (U.S. EPA, 1997a). The most common method of recycling metallic mercury is through

thermal treatment. Generally, the mercury-containing scrap is reduced in size and is heated in retorts or furnaces at about

538C (1000F) to vaporize the mercury. The mercury vapors are condensed by water-cooled condensers and collected

under water (Reisdorf and D'Orlando, 1984; U.S. EPA, 1984).

Vapors from the condenser, which may contain PM, organic compounds and possibly other volatile materials

from the scrap, are combined with vapors from the mercury collector line. This combined vapor stream is passed through

an aqueous scrubber to remove PM and acid gases (e.g., hydrogen chloride [HCl], SO ). From the aqueous scrubber, the 2

vapor stream passes through a charcoal filter to remove organic components prior to discharging into the atmosphere

(U.S. EPA, 1984).

The collected mercury is further purified by distillation and then transferred to the filling area. In the filling

area, special filling devices are used to bottle small quantities, usually 0.464 kg (1 lb) or 2.3 kg (5 lb) of distilled mercury. 

With these filling devices, the mercury flows by gravity through tubing from a holding tank into the flask until the flask

overflows into an overflow bottle. The desired amount of mercury is dispensed into the shipping bottle by opening a

valve at the bottom of the flask. The shipping bottle is then immediately capped after the filling and sent to the storage

area (Reisdorf and D'Orlando, 1984).

Chemical treatment can encompass several methods for aqueous mercury-containing waste streams. To

precipitate metallic mercury, the waste stream can be treated with sodium borohydride or the stream can be passed through

a zinc-dust bed. Mercuric sulfide can be precipitated from the waste streams by treatment with a water-soluble sulfide,

such as sodium sulfide. Ion-exchange systems can be used to recover ionic mercury for reuse, while mercuric ions can be

trapped by treatment with chemically modified cellulose (Cammarota, 1975).

4.2.5 Secondary Mercury Production

Secondary mercury production (mercury recycling) involves processing scrapped mercurycontaining products, industrial waste and scrap, and scrap mercury from government stocks. Secondary

mercury production is estimated to have accounted for approximately 0.4 Mg (0.4 tons) of mercury

emissions in 1995. Major sources of recycled mercury include dental amalgams, scrap mercury from

instrument and electrical manufacturers (lamps and switches), wastes and sludges from research

laboratories and electrolytic refining plants, and mercury batteries (U.S. EPA, 1997a). The recycling of

fluorescent lamps is discussed separately in Section 4.2.11.

There are two basic categories of secondary mercury production: recovery of liquid mercury

from dismantled equipment and mercury recovery from scrap products using extractive processes. On an

annual basis, the total quantity of mercury recovered as liquid mercury is much greater than that

recovered by extractive processes. Three areas have contributed to a large proportion of the liquid

mercury recovery category are: (1) dismantling of chlorine and caustic soda manufacturing facilities; (2)

recovery from mercury orifice meters used in natural gas pipelines; and (3) recovery from mercury

rectifiers and manometers. In each of these processes, the liquid mercury is drained from the dismantled

equipment into containers and sold on the secondary mercury market. The second category involves the

processing of scrapped mercury-containing products and industrial wastes and sludges using thermal or

chemical extractive processes because the mercury cannot be decanted or poured from the material. One

mercury recycler (Bethlehem Apparatus Company) estimated that this second category accounted for 15

to 20 percent of the total mercury reported as recycled from industrial scrap in 1995.
4-48

Table 4-11

1995 Major U.S. Mercury Recyclersa

Bethlehem Apparatus Company, Inc. Hellertown, PA

D. F. Goldsmith Chemical and Metals Corp. Evanston, IL

Mercury Refining Company, Inc. Albany, NY

 Plachy, 1997. a

In 1996, an estimated 446 Mg (492 tons) of mercury was recycled from industrial scrap.

According to the Mineral Industry Survey of Mercury, eight major companies were reported to be

involved in secondary mercury production using purchased scrap material (mercury recyclers) in 1996

(Plachy, 1997). The three dominate companies in this market are listed in Table 4-11. 

Information on specific emission control measures is very limited and site specific. If a scrubber

is used, mercury vapor or droplets in the exhaust gas may be recovered by condensation in the spray. 

There is no information to indicate that chemical filters would be effective in removing mercury vapors. 

No information was found for other control measures that are used in secondary mercury production

processes. Concentration in the workroom air due to mercury vapor emissions from the hot retort may

be reduced by the following methods: containment, local exhaust ventilation, dilution ventilation,

isolation, and/or personal protective equipment. No information was provided to indicate that these

systems are followed by any type of emission control device. Vapor emissions due to mercury transfer

during the distillation or filling stages may be reduced by containment, ventilation (local exhaust or

ventilation), or temperature control.

During production of mercury from waste materials using an extractive process, emissions may

vary considerably from one type of process to another. Emissions may potentially occur from the 

following sources: retort or furnace operations, distillation, and discharge to the atmosphere from the

charcoal filters. The major mercury emission sources are due to condenser exhaust and vapor emissions

that occur during unloading of the retort chamber. 

Mercury Refining Company reported results from two emission test studies conducted in 1994

and 1995 that showed average mercury emissions of 0.85 kg/Mg (1.7 lb/ton) of mercury recovered

(U.S.EPA, 1996b). In 1973, emission factors were estimated to be 20 kg (40 lb) per megagram (ton) of

mercury processed due to uncontrolled emissions over the entire process (Anderson, 1973). 

Mercury emission data were reported in the 1994 TRI only for Mercury Refining Company, Inc.,

in Albany, New York, and Bethlehem Apparatus Company in Hellertown, Pennsylvania. Mercury

Refining reported plant emissions to the atmosphere of 116 kg (255 lb) for 1994, and Bethlehem

Apparatus reported plant emissions to the atmosphere of 9 kg (20 lb) for 1994. The other major recycler,

D.F. Goldsmith, does not use extractive processes; their recycling is primarily from purchases of

mercury decanted from old equipment. Mercury emissions data were not available for the other five

facilities. 
4-49

To estimate mercury emissions from secondary mercury production, Bethlehem Apparatus and

Mercury Refining Company were assigned the emissions reported in the 1994 TRI and the remaining six

facilities were assigned the average of the emissions from the two reported facilities. The result is an

estimated 1994 total mercury emissions of 0.4 Mg (0.6 tons).

4.2.6 Electrical Apparatus Manufacturing

Mercury is one of the best electrical conductors among the metals and is used in five areas of

electrical apparatus manufacturing: electric switches, thermal sensing elements, tungsten bar sintering,

copper foil production, and fluorescent light production. Overall mercury emissions from electrical

apparatus manufacturing were estimated to be 0.31 Mg (0.34 ton) in 1995. No information on locations

of manufacturers of electrical apparatus that specifically contain mercury is available.

4.2.6.1 Electric Switches

The primary use of elemental mercury in electrical apparatus manufacturing is in the production

of electric switches (electric wall switches and electric switches for thermostats). Wall switches consist

of mercury, metal electrodes (contacts) and an insulator in button-shaped metal cans. Electric switches

containing mercury have been manufactured since the 1960s with approximately one million produced

annually.

The amount of mercury used for the manufacture of switches and thermostats decreased 50

percent from 155 tons in 1989 to 49 tons in 1996 (Plachy, 1997). This decrease in mercury use for the

manufacture of electric switches may be attributable to the shift to solid state devices and other

alternatives. The recent decrease in the construction of houses may have also contributed to the decrease

in mercury use for electric switch manufacture (Cole et al., 1992).

The amount of mercury disposed each year in electric switches compared to the amount of

mercury in electric switches in use is small. One recent study estimated that 10 percent of switches are

discarded after 10 years, 40 percent after 30 years and the remaining 50 percent after 50 years (U.S.

EPA, 1992a). Average unit life for mercury thermostats exceeds 20 years, with upgrading, remodeling

or building demolition being the principal causes for removal from service (National Electrical

Manufacturers Association, 1995). In addition, a few will be discarded due to leakage or some other

failure.

Table 4-12 summarizes the discards of mercury in electric switches. In these estimates it was

assumed that there is no recycling of mercury from discarded switches. In 1994, however, Honeywell,

Inc., a major manufacturer of thermostats announced a pilot project in Minnesota to recycle mercury

thermostats. Homeowners and contractors can send unneeded thermostats back to Honeywell so the

mercury can be removed and recycled. In addition, in 1995, U.S. EPA announced a "Universal Waste

Rule" (which includes thermostats) that effectively allows for the transportation of small quantities of

mercury from specific products. This ruling is intended to encourage recycling. Until programs such as

these are fully implemented, it is unclear how much the mercury discards from this type of product will

decline in MSW.
4-50

Electric Switch Manufacturing Process

The wall switches are manufactured by first assembling a component consisting of a metal ring, a

glass preform, a ceramic center, and a center contact. This subassembly is then transferred to a rotating

multistation welding machine, located in an isolation room, where it is filled with approximately 3 g

(0.11 oz.) of mercury. The filled subassembly is placed in the button-shaped can, evacuated, and welded

shut. The assembled buttons then leave the isolation room and are cleaned, zinc-plated and assembled with

other components to form the completed wall switches (Reisdorf and D'Orlando, 1984).

Thermostat switches are constructed using a short glass tube with wire contacts sealed in one end of

the tube. First, metal electrodes (contacts) are inserted into small tubes. The tubes are then heated at one

end, constricted and crimped closed around the electrodes (sealing the electrodes into the glass tube), and the

apparatus is cleaned. The subassembly is then transferred to the isolation fill room where mercury is added. 

The open end of the mercury-filled tube is then heated, constricted and sealed. The filled tubes then leave the

isolation room, and wire leads are attached to the electrode contacts, which completes the switch assembly

(Reisdorf and D'Orlando, 1984).

During electric switch manufacture, mercury may be emitted during welding or filling operations, as

a result of spills or breakage, during product testing, and as a result of product transfer. Often, emissions can

be controlled by using effective gaskets and seals to contain mercury in the process streams. Also, good

work practices, such as discarding rejected and broken switches under water and reducing the temperature in

the fill room, can effectively suppress mercury vaporization. Furthermore, local exhaust ventilation, customdesigned to fit specific equipment, can reduce mercury vapor and mercury PM (Reisdorf and D'Orlando,

1984).

4.2.6.2 Thermal Sensing Instruments and Tungsten Bar Sintering

A thermal sensing instrument consists of a temperature-sensing bulb, a capillary tube, a mercury

reservoir and a spring-loaded piston. The bulbs are made by cutting metal tubing to the correct size,

welding a plug to one end of the tube and attaching a coupling piece to the other end. A capillary is cut

to a specified length and welded to the coupling at the open end of the bulb. The other end of the

capillary is welded to a "head" that houses the mechanical section of the sensor. The bulb and capillary

assembly are filled with mercury by a multistation mercury filling machine that is housed in a ventilated

enclosure. After filling, the sensor is transferred to a final assembly station, where a return spring and

plunger are set into a temporary housing on the head of the sensor. In order to complete the temperature

instrument, the sensor is then attached to a controller and/or indicating device (Reisdorf and D'Orlando,

1984).

Mercury is also used in tungsten bar sintering. Tungsten is used as a raw material in

manufacturing incandescent lamp filaments. The manufacturing process starts with tungsten powder

pressed into long, thin bars of a specified weight. These bars are presintered and then sintered using a

high-amperage electrical current. During the tungsten bar sintering process, mercury is used as a

continuous electrical contact. The mercury contact is contained in pools (mercury cups) located inside

the sintering unit.

After the sintering process is completed, the bars are cooled to ambient temperature to determine

the density of the tungsten bar. Metallic mercury is normally used in these measurements because of its

high specific gravity. In order to calculate the density of the tungsten bar, the tungsten 
4-51

Table 4-12

Discards of Mercury in Electric Switchesa

Year

Electric Switch Production Weight of Mercury in

Switches (tons)

Weight of Mercury Discarded in

MSW (tons)

1987 1,000,000 3.9 0.39

1988 1,000,000 3.9 0.39

1989 1,000,000 3.9 0.39

1995 1,000,000 3.9 1.93

2000 1,000,000 3.9 1.93

 U.S. EPA, 1992a. a

bars are dipped into a pool of mercury and the weight of the displaced mercury is determined. When the

bar is removed from the mercury pool, the mercury is brushed off into a tray of water that is placed in

front of the pool (Reisdorf and D'Orlando, 1984).

No specific information on emission control measures for thermal sensing elements and tungsten

bar sintering was found in the literature. It is assumed that mercury is emitted during the filling process

for thermal sensing elements and during sintering and final density measurements for tungsten bar

sintering (U.S. EPA, 1997a).

4.2.6.3 Copper Foil Production

High-purity copper foil, used as a laminate in printed circuit boards, is produced by an

electrodeposition process using mercury as the electrical contacts. The initial step in the foil production

process is the dissolution of scrap copper in sulfuric acid to form copper sulfate. The solution is then fed

to the plating operation, where the copper ions are electrodeposited on rotating drums as copper metal. 

During the electrodeposition process, a current passes between a lead anode and a rotating drum cathode. 

As the drum rotates, the copper metal is electrodeposited on the drum surface in the form of a continuous

thin foil sheet. The rotating drum requires using a rotating electrical contact between the electrical

connection and the drum surface. Elemental mercury is used as the continuous contact between the

rotating shaft of the drum and the electric connections. The liquid mercury is contained in a well located

at one end of the rotating drum shaft (Reisdorf and D'Orlando, 1984).

During copper foil production, mercury can be emitted from the drum room and the treatment

room of the copper plating process. Ventilated enclosures, with exhaust gases directed to mercury vapor

filters, can be used to control mercury emissions, as can reducing the temperature of the mercury wells

(Reisdorf and D'Orlando, 1984).
4-52

4.2.6.4 Fluorescent Lamps

All fluorescent lamps contain elemental mercury as mercury vapor inside the glass tube. 

Mercury has a unique combination of properties that make it the most efficient material for use in

fluorescent lamps. Of the 500-600 million mercury-containing lamps sold in the United States annually,

approximately 96 percent are fluorescent lamps. It is estimated in that approximately the same amount

of lamps are disposed of on an annual basis (National Electrical Manufacturers Association, 1992). In

fluorescent lamp production, precut glass bulbs are washed, dried and coated with a liquid phosphor

emulsion that deposits a film on the inside of the lamp bulb. Mount assemblies are fused to each end of

the glass lamp bulb, which is then transferred to an exhaust machine. On the exhaust machine, the glass

bulb is exhausted and 15 to 250 mg (3.3 x 10 to 5.5 x 10 lb) of mercury is added. Some of the -5 -4

mercury combines with the emulsion on the interior of the bulb and remains there over the life of the

bulb. The glass bulb is filled with an inert gas and sealed. After the lamp bulbs are sealed, metal bases

are attached to the ends and are cemented in place by heating.

The names and division headquarters of the fluorescent lamp manufacturers in the United States

in 1995 are shown in Table 4-13 (U.S. EPA, 1997a).

Table 4-13

1995 U.S. Fluorescent Lamp Manufacturers' Headquarters a

Company Division headquarters

Duro-Test Corp. North Bergen, NJ

General Electric Cleveland, OH

OSRAM Corp. Montgomery, NY b

Philips Lighting Company Somerset, NJ

 U.S. EPA, 1997a. a

 National Electrical Manufacturers Association, 1995. b

During fluorescent lamp manufacturing, mercury can be emitted by transfer and parts repair

during mercury handling; by the mercury injection operation; and from broken lamps, spills and waste

material. Mercury air levels during lamp production steps are reduced by process modifications,

containment, ventilated enclosures, local exhaust ventilation, and temperature control (Reisdorf and

D'Orlando, 1984).

4.2.6.5 Emissions Summary for Electrical Apparatus Manufacturing

While mercury may be emitted from all of the aforementioned areas of electrical apparatus

manufacturing, no specific data for mercury emissions from these areas were found in the literature and

no emission test data were available to calculate mercury emissions from each area. One 1973 U.S. EPA

report presents an emission factor of 4 kg of mercury emitted for each megagram of mercury used

(8 lb/ton) in overall electrical apparatus manufacture (Anderson, 1973). This factor only pertains to

emissions generated at the point of manufacture. This emission factor should be used with extreme

caution, however, as it was based on engineering judgment and not on actual test data and because

production and mercury control methods have probably changed considerably since 1973 to prevent

waste and limit worker exposure. The emission factor may, therefore, substantially overestimate

mercury emissions from this source.
4-53

In 1996, 78 Mg (86 tons) of mercury were used in all electrical apparatus production (29 Mg

[32 tons] for electric lighting and 49 Mg [54 tons] for wiring devices and switches) (Plachy, 1997).

Multiplying the emission factor above by the 1992 usage gives a mercury emission estimate of 0.31 Mg

(0.34 ton) for electrical apparatus manufacture. Because of the lack of reliability of the emission factor,

a high degree of uncertainty is associated with this emission estimate.

4.2.7 Carbon Black Production

The majority of U.S. manufactured carbon black (over 98 percent) is produced using a highly

aromatic petrochemical or carbochemical heavy oil feedstock containing mercury. In 1995, mercury

emissions from carbon black production were estimated to be 0.25 Mg (0.28 ton). This estimate is

expected to be an overestimate because it is based on production capacity and not on actual production. 

Table 4-14 lists the names, locations and annual capacities of U.S. producers of carbon black in 1995

(SRI International, 1996). The geographic distribution of these facilities is shown in Figure 4-13.

High-performance fabric filters are reported to be used to control PM emissions from main

process streams during the manufacture of carbon black. The fabric filters can reduce PM emissions to

levels as low as 6 milligrams per normal cubic meter (mg/Nm ) (0.003 gr/dscf). Mercury emissions from 3

the reactor are primarily in the vapor phase, and these emissions will proceed through the main process

streams to the fabric filters as a vapor. If the mercury remains in the vapor phase, the mercury control

efficiency of the fabric filters is expected to be low. If the product gas stream is cooled to below 170C

(325F), the fabric filter may capture a significant fraction of the condensed mercury, thus providing

some degree of emission control (Taylor, 1992).

Mercury, which is present in the oil feedstock, can be emitted during the pyrolysis step. No data

are available, however, on the performance of the fabric filter control systems for mercury emissions. 

The only available data are for emissions from the oil-furnace process. These data show mercury

emission to be 1.5 x 10-4 kg/Mg (3 x 10 lb/ton) from the main process vent (Serth and Hughes, 1980). -4

The source of these data could not be obtained in order to validate the emission factors. Because the

factors are not verified, they are considered to be of limited reliability.

In 1995, the total capacity for carbon black production was 1.66 x 10 Mg (1.83 x 10 tons) (SRI 6 6

International, 1996). Multiplying the total capacity by the emission factor above gives a mercury

emission estimate of 0.25 Mg (0.28 tons). This estimate may be greater than the actual emissions

estimate because it is based on production capacity and not on actual production. On the other hand, this

estimate may understate the actual mercury emissions because the data are from the oil-furnace process

only and not the main process streams.
4-54

Table 4-14

1992 U.S. Carbon Black Production Facilitiesa

Company Location Type of

processb

Annual capacityc

10 Mg 3 10 tons 3

Cabot Corporation, North

American Rubber Black

Division

Franklin, Louisiana F 141 178

Pampa, Texas F 29 33

Villa Platte, Louisiana F 100 110

Waverly, West Virginia F 91 100

Chevron Corporation,

Chevron Chemical Company,

subsidiary, Olevins and

Derivatives Division

Cedar Bayou, Texas A 9 10

Degussa Corporation Arkansas Pass, Texas F 54 60

Belpre, Ohio F 54 60

New Iberia, Louisiana F 109 120

Ebonex Corporation Melvindale, Michigan C 4 4

Engineered Carbons, Inc. Baytown, Texas F 86 95

Borger, Texas F and T 102 112

Orange, Texas F 61 67.5

General Carbon Company Los Angeles, California C 0.5 0.5

Hoover Color Corporation Hiwassee, Virginia C 0.5 0.5

Phelps Dodge Corporation

Colombian Chemical Company,

subsidiary

El Dorado, Arkansas F 57 63

Moundsville, West Virginia F 88 98

North Bend, Louisiana F 100 110

Ulysses, Kansas F 36 40

Sir Richardson Carbon

Company

Addis, Louisiana F 120 133

Big Spring, Texas F 54 60

Borger, Texas F 129 143

Witco Corporation

Continental Carbon Company,

subsidiary

Phenix City, Alabama F 36 40

Ponca City, Oklahoma F 120 133

Sunray, Texas F 59 65

TOTAL 1,660 1,830

 SRI International, 1996. a

 A = acetylene decomposition; F = furnace; C = combustion; T = thermal. b

 Capacities are variable and based on SRI estimates as of January 1, 1996. c
Figure 4-13

Carbon Black Manufacturing Facilities

The Carbon Black Production Process

Three primary raw materials used in the production of carbon black are preheated

feedstock (either the petrochemical oil or carbochemical oil), which is preheated to a temperature

between 150 and 250°C (300 and 480”F), preheated air and an auxiliary fuel such as natural gas.

A turbulent, high-temperature zone is created in the reactor by combusting the auxiliary fuel, and

the preheated oil feedstock is introduced in this zone as an atomized spray. In this zone of the

reactor, most of the oxygen is used to bum the auxiliary fuel, resulting in insufficient oxygen to

combust the oil feedstock. Thus, pyrolysis of the feedstock is achieved, and carbon black is

produced. Most of the mercury present in the feedstock is emitted in the hot exhaust gas from the

reactor (Taylor, 1992; Yen, 1975).

The product stream from the reactor is quenched with water, and any residual heat in the

product stream is used to preheat the oil feedstock and combustion air before the carbon is

recovered in a fabric filter. Carbon recovered in the fabric filter is in a fluffy form. The fluffy

carbon black may be ground in a grinder, if desired. Depending on the end use, carbon black

may be shipped in fluffy form or in the form of pellets. Pelletizing is done by a wet process in

which carbon black is mixed with water along with a binder and fed into a pelletizer. The pellets

are subsequently dried and bagged prior to shipping (Taylor, 1992; Yen, 1975).

4-55
4-56

4.2.8 Lime Manufacturing

Lime is produced in various forms, with the bulk of production yielding either hydrated lime or

quicklime. In 1994, producers sold or used 17.4 x 10 Mg (19.2 x 10 tons) of lime produced at 6 6

109 plants in 33 States and Puerto Rico. The 1994 production represented a 3.6 percent increase over

1993 production. The leading domestic uses for lime include steelmaking, flue gas desulfurization, pulp

and paper manufacturing, water purification, and soil stabilization (Miller, 1996). Total mercury

emissions from lime manufacturing are estimated to be 0.1 Mg (0.1 tons) per year.

Table 4-15 identifies the top 10 lime-producing plants in the United States, in order of total

output for 1994 (Miller, 1996). Lime production is geographically concentrated as demonstrated by

1989 production data, when 63 percent of the U.S. total was produced in seven States (in order of

decreasing production: Missouri, Ohio, Pennsylvania, Alabama, Kentucky, Texas and Illinois) (Bureau

of Mines, 1991).

Fuels, including primarily coal, oil, petroleum coke, or natural gas, are used to provide the

energy for calcination. Petroleum coke is usually used in combination with coal. Auxiliary fuels may

include shredded municipal garbage, chipped rubber, or waste solvent. Mercury is expected to be

present in the coal, oil, and possibly in appreciable quantities in any waste-derived fuels. Any mercury

emitted from fuel combustion will occur during the calcination step and will be discharged as vapor kiln

exhausts.

The quicklime that is produced by calcination can be hydrated with water to produce hydrated

lime or slaked lime (Ca(OH) ). The hydration step may be immediately preceded by some crushing, 2

pulverizing and separation of dolomitic quicklime to form high calcium and dolomitic quicklime. These

processes and handling, storage and transfer are not likely sources for mercury emissions during lime

production.

Air pollution control devices for lime kilns are primarily used to recover product or control

fugitive dust and PM emissions. Calcination kiln exhaust is typically routed to a cyclone for product

recovery and then routed through a fabric filter or ESPs to collect fine particulate emissions. Other

emission controls found at lime kilns include wet scrubbers (typically venturi scrubbers). How well

these various air pollution control devices perform relative to vapor phase mercury emissions in lime

production is not well documented. The control efficiencies are expected to be similar to those observed

in the production of portland cement, however, because of the similarities in the process and control

devices.
4-57

Table 4-15

Lime Producers in the U.S. in 1994

No. of Lime production x 10 Mg (x 10 tons) 3 3

State plants Hydrateda Quicklimea Totala

Alabama 4 184 (203) 1,470

(1,620)

1,660 (1,829)

Arizona, Nevada, Utah 8 243 (268) 1,570

(1,730)

1,810 (1,995)

California 7 26 (29) 178 (196) 203 (224)

Colorado, Montana,

Wyoming

10 -- (--) 335 (369) 335 (369)

Idaho, Oregon, Washington 8 25 (28) 597 (658) 622 (685)

Illinois, Indiana, Missouri 8 464 (511) 2,910

(3,207)

3,380 (3,725)

Iowa, Nebraska, South

Dakota 

5 W (W) b W (W) (242) (267) c

Kentucky, Tennessee, West

Virginia

5 132 (145) 1,800

(1,984)

1,930 (2,127)

Michigan 9 26 (29) 611 (673) 637 (702)

North Dakota 3 -- (--) 108 (119) 108 (119)

Ohio 9 W (W) W (W) (1,850) (2.039) c

Pennsylvania 8 263 (290) 1,330

(1,466)

1,590 (1,752)

Puerto Rico 1 23 (25) <0.5 (<0.6) 23 (25)

Texas 6 471 (519) 740 (815) 1,210 (1,333)

Virginia 5 121 (133) 621 (684) 742 (818)

Wisconsin 4 124 (137) 383 (422) 507 (559)

Otherd 9 213 (235) 2,430

(2,678)

2,640 (2,909)

Total 109 2,310

(2,546)

15,100

(16,640)

17,400 (19,175)

Source: Miller, 1996.

 Metric ton data rounded by the U.S.G.S. to three significant digits; may not add to totals shown. a

 Witheld to avoid disclosing company proprietary data; included in Othercategory. b

 Total included in total for Othercategory. c

 Includes Arkansas, Louisiana, Massachusetts, Minnesota, Oklahoma, and data indicated by W. d
4-58

Table 4-16

1994 U.S. Primary Lead Smelters and Refineriesa

Smelter Refinery 1994 Lead Production

Tons (Megagrams)

ASARCO, East Helena, MT ASARCO, Omaha, NEb 65,800 (72,500)

ASARCO, Glover, MO ASARCO, Glover 125,000 (137,800)

Doe Run (formerly St. Joe) Doe Run, Herculaneum, MO 200,000 (220,400)

 Source: Smith, 1996. a

 Closed permanently for lead refining as of May 31, 1996. There is limited refinery capacity at East Helena, MT. b

4.2.9 Primary Lead Smelting

Primary lead smelters recover lead from a sulfide ore, which may contain mercury. The smelters

emitted an estimated 0.10 Mg (0.11 tons) of mercury into the atmosphere in 1994. Table 4-16 lists the

locations and 1994 production rates of the two primary lead smelters that are currently operating in the

United States; the locations of these smelters are displayed in Figure 4-14.

Primary lead smelters use high-efficiency emission control systems to reduce the levels of PM

and SO from the blast furnace and sintering machines. Centrifugal collectors (cyclones) are 2

used in conjunction with baghouses or ESPs for PM control. Control of SO emissions from sintering is 2

achieved by absorption to form sulfuric acid in the sulfuric acid plants, which are commonly part of lead

smelting plants. Because mercury is emitted from these as a vapor and these PM control systems often

operate at temperatures at which mercury has a significant vapor pressure, these PM control devices are

expected to have little effect on mercury emissions from the sintering machine and blast furnace. In

contrast, sulfuric acid plants are expected to be relatively well controlled for mercury because of the low

temperatures and high particulate removal efficiency of the APC device. No data are available, however,

on performance of these systems with respect to mercury emissions (U.S. EPA, 1988).

Mercury, which may be present in the ore, may be emitted during the sintering and blast furnace

steps and in the drossing area because these processes take place at high temperatures.

No recent mercury emission factors are available for the two currently operating primary lead

smelters; none of the three primary lead smelters reported mercury emission data in the 1994 TRI. The

only available mercury emission factors were provided by industry for a custom smelter operated by

ASARCO in El Paso, Texas which ceased operating in 1985 (Richardson, 1993). Because the El Paso

facility data were based on ores with a variable mercury content, and the current major sources of lead

ore have a very low mercury content, use of those emission factors will lead to an overestimation of 

current emissions. A better estimating method is to use the actual mercury content of the ore and 

emissions based on those data. The major domestic source of lead ore concentrate is from the southeast

Missouri area near the Glover and Herculaneum smelters. Data on mercury content estimate in lead
Figure 4-14

Primary Lead Smelters

The Primary Lead Smelting Process

Recovery of lead from the lead ore in primary lead smelters consists of three main steps: sintering, reduction

and refining. The sintering machine, which converts lead sulfide in the ore to lead and lead oxide, is a continuous steel

pallet conveyor belt. Each pallet consists of perforated grates, beneath which are wind boxes connected to fans to

provide a draft through the moving sinter charge. The sintering reactions on the grate take place at about 1000°C

(1832°F). Because mercury and its compounds volatilize below this temperature, most of the mercury present in the

ore is emitted as a vapor in the sintering machine exhaust gas as elemental mercury or as mercuric oxide.

Reduction of the sintered lead is carried out in a blast furnace at a temperature of 1600°C (2920°F). The

furnace is charged with a mixture of sinter (80 to 90 percent of charge), metallurgical coke (8 to 14 percent of charge)

and other materials, such as limestone, silica, litharge, and other slag-forming constituents. In the blast furnace, the

lead sulfate and lead oxide in sinter is reduced to lead. The heat for the reaction is supplied by the combustion of coke.

Impurities are removed from the furnace as slag, which is either processed at the smelter for its metal content, shipped

to treatment facilities, or landfilled. The impurities include arsenic, antimony, copper, and metal sulfides and silicates.

Lead bullion, which is the primary product, undergoes a preliminary treatment to remove impurities, such as copper,

sulfur, arsenic, antimony, and nickel, before carrying out further refining. Any residual mercury left in the ore after

sintering will be emitted during the reduction step (U.S. EPA, 1988).

The lead bullion is refined in cast iron kettles. Refined lead, which is 99.99 to 99.999 percent pure is cast

into pigs for shipment (U.S. EPA, 1988). Mercury emissions from refining operations are expected to be negligible.

4-59
4-60

concentrates from this area indicate the mercury concentration to be less than 0.2 ppm (Richardson,

1993). Based on this concentration, the mercury content is estimated to be 0.4 x 10 pounds of mercury -3

per ton of ore concentrate. Particulate matter (PM) emission factors were used with a mercury

concentration of 0.2 ppm to estimate 1994 mercury emissions. The estimated 1994 lead in ore

concentrate quantity was 3.7 x 10 Mg (4.07 x 10 tons) (Smith, 1996). Based on background 5 5

information in the NSPS for lead smelters, 100 units of ore yields 10 units of ore concentrate, 9 units of

sinter, and 4.5 units of refined lead (EPA, 1974). The following PM emission factors from AP-42 (EPA,

1995b) were used for 3 emission sources in the process:

sinter machine (weak gas): 0.051 kg/Mg (0.10 lb/ton) of sinter produced

sinter building fugitives: 0.118 kg/Mg (0.24 lb/ton) of sinter produced

blast furnace: 0.21 kg/Mg (0.43 lb/ton) of bullion

Combining these PM figures with the mercury content and ore fractionation figures above to calculate

emissions from these 3 processes, the upper limit for total mercury emissions from primary lead smelting

was estimated to be 0.10 Mg (0.11 tons) per year.

4.2.10 Primary Copper Smelting

Copper is recovered from a sulfide ore principally by pyrometallurgical smelting methods. The

ore contains significant quantities of arsenic, cadmium, lead, antimony and mercury. Table 4-17 gives

the locations and 1996 production capacities of primary copper smelters currently operating in the United

States; these smelter locations are displayed in Figure 4-15.

Copper smelters use high efficiency air pollution control options to control PM and SO2

emissions from smelting furnaces and convertors. Electrostatic precipitators are the most common PM

control device at copper smelters. Control of SO emissions is achieved by absorption to sulfuric acid in 2

the sulfuric acid plants, which are common to all copper smelters.

A recent analysis of the seven copper smelters currently operating in the U.S. has been

performed. Mercury emission rates from these seven smelters are presented in Table 4-18 along with the

mercury concentration of ore. These data, self-reported by industry, show that emissions range from less

than 1 lb/year to 40 lbs/year. These emission rates are based on both stack testing and engineering

judgment. As a result, the U.S. EPA estimates 1994 nationwide mercury emissions from primary copper

smelters to be about 0.06 Mg/year (0.06 tons/year).

4-62

Table 4-18

Mercury Ore Concentrate and Emissions from

Primary Copper Smelters in the U.S.

Smelter Mercury in Ore Mercury Basis of

Concentrate Emissions Emission Values

lb/yr lb/yr

ASARCO - El Paso 1,769 1.8 Emissions Test

ASARCO - Hayden 2,444 35 Emissions Test and

Engineering Judgment

Copper Range 940 1,951 Emissions Test b

Cyprus Miami CBI 34 Emissions Test a

Kennecott NA 35 Emissions Test and a

Engineering Judgment

BHP Copper Co. 2,240 40 Emissions Test and

Engineering Judgment

Phelps Dodge-Hidalgo 5,768 0.09 Engineering Judgment

Phelps Dodge-Chino 585 7.5 Engineering Judgment

 CBI means Confidential Business Information that is unavailable to the public. NA means not available. a

Ceased operation in February 1995. b 
4-63

The Primary Copper Smelting Process

The copper smelting process sequentially involves drying ore concentrates, smelting of ore concentrates to

produce matte, converting matte to produce blister copper, and fire refining the blister copper in an anode furnace. After

fire refining, the 99.5 percent pure copper is cast into "anodes" and sent to an electrolytic refinery for further impurity

removal (Buonicore and Davis, 1992).

All of the currently operating copper smelters use either fluid bed or rotary kiln dryers to dry the concentrate. 

Temperatures in the dryer are not high enough to vaporize any mercury in the ore concentrate. Roasting of ores is no

longer used because the off gases from the roasting process were too low in SO to be processed in the sulfuric acid plant. 2

Smelting produces a copper matte by melting the hot ore concentrates with siliceous flux in a furnace. The

mattes produced by domestic smelters range from 35 to 65 percent copper. Smelting furnace technologies operate at

temperatures well above the boiling point of mercury with operating ranges as high as 2500C (4530F). Any mercury

contained in the concentrate will likely be emitted during the flash smelter process step and directed to the sulfuric acid

plant (Buonicore and Davis, 1992). The gas stream to the sulfuric acid plant passes through three to five control devices,

such as dry ESPs, cyclones, scrubbing towers, cooling towers and acid mist ESPs. These control devices are required to

remove metal impurities to prevent destruction of the catalyst in the acid plant. Any mercury volatilizing in the smelting

furnace is removed in these multistage control systems and in the sulfuric acid plant. Limited data on sulfuric acid plant

sludges show that the mercury is present in measurable concentrations. This mercury is recycled back to the flash

converter and vaporized again into the control system. This appears to set up an internal recycling loop for the mercury,

which is ultimately discarded with the solid waste.

The final step in the production of molten "blister" copper is converting. Converting eliminates remaining iron

and sulfur impurities, leaving 98.5 to 99.5 percent pure copper. Converting involves molten matte, siliceous flux and

scrap copper being charged in a rotating cylindrical shell, where air or oxygen rich air is blown through the molten matte. 

Blowing and slag skimming are repeated until relatively pure Cu S, called "white metal" accumulates in the bottom of the 2

converter. A renewed air blast then oxidizes the copper sulfide to SO , leaving blister copper. Blister copper is then 2

removed and transferred to refining facilities. Further purification may involve fire refining and electrolytic refining

(Buonicore and Davis, 1992).

4.2.11 Fluorescent Lamp Recycling

In order to reduce the net amount of mercury released to the environment, recycling of

fluorescent lamps has become a more common practice. The recycling process begins with the crushing

of the lamps to extract the white phosphor powder in them, which contains the bulk of mercury in lamps. 

Lamps can be crushed either by a mobile crushing unit at the point of collection, or by a centralized

stationary crushing unit. Mercury emissions from crushing operations may be reduced using a vacuum

collection system. In a vacuum collection system, air is passed through a cyclone to remove glass

particles, followed by a filter to remove the phosphor powder, and a carbon adsorber to capture the

mercury vapor, before being exhausted (Battye et al., 1994).

Mercury is recovered from crushed lamps by heating the crushed material to vaporize the

mercury and then cooling the off gas stream to condense liquid elemental mercury (Battye et al., 1994). 

This can be accomplished in closed vessels called retorts or in open-hearth furnaces, ovens, or rotary

kilns referred to as roasters. Retorting generally gives higher recovery rates than does roasting and is

well suited to wastes containing volatile forms of mercury (Battye et al., 1994).
4-64

Because fluorescent lamp recycling and lamp breakage are considered separate source categories

in this study, it is difficult to categorize facilities which perform only the crushing operation and send the

recovered powders to other facilities to perform the mercury extraction. According to industry sources,

this difficulty is compounded by the fact that many of the lamp crushing facilities deal not only with

lamp bulbs but also other types of mercury scrap. There are approximately six or seven such sites in

Florida, seven in Ohio, three or four each in California, Wisconsin and Minnesota, and some in

Louisiana, New York, and Texas (Lawrence, 1997).

As presented previously in Figure 3-1, 2 percent of fluorescent lamps are estimated to be

recycled each year. Industry estimates that 75 million lamps will be recycled in 1997, representing 12.5

percent of the 500-600 million lamps which are disposed (O’Connell, 1997). Air emission and mass

balance information for fluorescent lamp recycling facilities was only available from one company. 

Based on this information, it was determined that only 1 percent of the mercury entering the recycling

facility is emitted. This is equal to 0.005 Mg, or 0.02 percent of the mercury entering the MSW system

(Truesdale, 1993). 

4.2.12 Battery Production

Historically, mercury has been used in batteries for two purposes. The first use is as a

component in the zinc-mercury amalgam used as the anode in mercury oxide (also known as mercuryzinc) and alkaline batteries and as a component in the cathode of mercury oxide batteries. The second

use was to inhibit side reactions and corrosion of the battery casing material in carbon-zinc and alkaline

batteries. Prior to the late 1980s, most primary batteries and some storage batteries contained mercury in

the form of mercuric oxide (HgO), zinc amalgam (Zn-Hg), mercuric chloride (HgCl ), or mercurous 2

chloride (Hg Cl ) (White and Jackson, 1993). As a result of technological improvements made by the 2 2

battery industry, the use of mercury is being phased out of battery production. From 1989 to 1992, the

use of mercury in battery production decreased 94 percent (Bureau of Mines, 1992). Because only one

type of battery, mercuric oxide batteries, still used mercury to any measurable degree as of the end of

1992, it is the only battery discussed in this section. In 1992, an estimated 0.02 Mg (0.02 ton) of

mercury was emitted from the production of batteries. Table 4-19 lists the manufacturers of mercuric

oxide, alkaline manganese and zinc-carbon batteries and the associated emissions reported in the 1990

TRI (U.S. EPA, 1992e). The TRI does not distinguish the type of battery each facility produces.

Mercuric oxide batteries fall into two categories: button cells and larger sizes. Most mercuric

oxide batteries sold for personal use are button cells. Button cells are small, circular, relatively flat

batteries that are used in transistorized equipment, walkie-talkie's, hearing aids, electronic watches, and

other items requiring small batteries. Mercuric oxide batteries are widely used for applications that

require reliability and a constant rate of discharge, including medical and military applications. Larger

mercuric oxide batteries, which often resemble 9-volt or fat AA batteries in size or shape, are produced

for a variety of medical, industrial, military, and other non-household devices (Dierlich, 1994). The

mercury content in mercuric oxide batteries is typically 33 percent to 50 percent mercury by weight and

cannot be reduced without proportionally reducing the energy content of these batteries. Acceptable

alternative batteries are available for almost all applications of household mercuric oxide batteries (Cole

et al., 1992; Balfour, 1992).
4-65

Table 4-19

1992 U.S. Mercuric Oxide, Alkaline Manganese, or 

Zinc-Carbon Button Cell Battery Manufacturers a

Manufacturer Production site

1990 Mercury TRI emissions

kg (lb)b

Alexander Manufacturing Company

(AMC, Inc.)

Mason City, IA 0 (0)

Duracell, USA Cleveland, TN

LaGrange, GA

Lancaster, SC

Lexington, NC

NRc

NR

9 (20)

3 (70)

Eagle-Picher Industries, Inc. Colorado Springs, CO NR

Eveready Battery Company, Inc. Maryville, MO

Red Oak, IA

Fremont, OH

Bennington, VT

Asheboro, NC (2 plants)

14 (30)

NR

NR

1 (2)

2 (5)

Mutecd Columbus, GA (Corporate offices) NR

Rayovac Corp. Madison, WI

Fennimore, WI

Portage, WI

0 (0)

5 (10)

NR

 U.S. EPA, 1993a. a

 U.S. EPA, 1992e. b

 NR = Not reported, company did not report mercury emissions in 1990 TRI. c

 Mutec is a joint venture between Eastman Kodak and Panasonic. d

Mercuric oxide-zinc cells use mercuric oxide (mixed with graphite and manganese dioxide) as

the cathode and a zinc amalgam at the anode. In producing the cathodes, granulated mercuric oxide,

manganese dioxide, and granulated graphite are manually metered through a hopper to the blending area

(U.S. EPA, 1984). This mixture is then pelletized in a rotary press. The pellets are consolidated into

plastic trays and are then sent to the production lines for cell assembly. For the production of the anodes,

elemental mercury and zinc powder are blended along with electrolyte and a binder to produce an anode

gel (Rauh, 1991). The completed anodes and cathodes are then sent to the cell manufacturing area. 

Separators, electrolytes and other components are assembled with the anode and cathode to produce the

HgO-Zn cell. Assembly may be automatic or semiautomatic. The assembled cathode, anode,

electrolyte, and cover are sealed with a crimper.

During the manufacture of mercuric oxide batteries, mercury may be emitted from grinding,

mixing, sieving, pelletizing, and/or consolidating operations as PM and as vapor emissions. Baghouses

are used to control PM emissions from the mixing/blending and processing steps in the production of

cathodes. Mercury vapor emissions from the anode processing and cell manufacturing areas are

generally discharged to the atmosphere uncontrolled. Ventilation air in the assembly room is
4-66

recirculated through PM filters. One plant reported an average of 73 percent mercury vapor removal

efficiency in the cell assembly room when an air handler system, consisting of a PM prefilter and a

charcoal filter, was operated using 75 percent recirculating air and 25 percent fresh air (Reisdorf and

D'Orlando, 1984).

The only reported emission factor for a mercuric oxide production facility was for one plant in

Wisconsin (Bureau of Air Management, 1986). This facility used a combination of a baghouse and

charcoal filter to treat the exhaust ventilation air. Annual use of mercury was 36.07 Mg (39.8 tons), and

annual emissions were reported as 36.3 kg (80 lb) of mercury as HgO particles. The mercury emission

factor for battery manufacture based on these data is 1.0 kg/Mg (2.0 lb/ton) of mercury used.

Several factors limit the reliability of this emission factor. First, the facility no longer produces

mercuric oxide batteries. The processes and emission controls may be substantially different for existing

mercuric oxide facilities, although no information on different process or controls was provided to U.S.

EPA from one current manufacturer. Second, no information is presented on the bases of the emission

factor, but the mercury emission quantity is presumed to be an engineering estimate by the manufacturer

because no reference is made to any emissions testing performed at the facility. Finally, this factor is

based on only one specific site, and that facility may not represent all mercuric oxide battery

manufacturing facilities.

Emission source data from a study of an integrated mercury button cell plant are summarized in

Table 4-20 (U.S. EPA, 1984). Major emission points were the pelletizing and consolidating operations

(up to 42.46 g/d [0.094 lb/d]) and cell assembly (28.58 g/d [0.063 lb/d]). Emission controls were not in

place for mercury vapor emissions from the main plant (U.S. EPA, 1984). This plant reported total

mercury emissions of 3.2 kg (7 lb) in the 1990 TRI (U.S. EPA, 1992e).

In 1995, less than 0.5 Mg (<0.6 tons) of mercury were used in the production of batteries in the

United States (Plachy, 1996). Multiplying the mercury usage by the emission factor developed for the

facility in Wisconsin gives a mercury emission estimate of 0.0005 Mg (0.0006 tons) for 1995. This

estimate is highly uncertain, however, because of the concerns discussed above about the reliability of

the emission factors (U.S. EPA, 1997a). Mercury emissions to the atmosphere when batteries are

disposed are accounted for in the emission estimate for MWCs and MWIs, as discussed in Section 4.1 of

this Report.
4-67

Table 4-20

Emission Source Parameters for an Integrated

Mercury Button Cell Manufacturing Facilitya

Building/source descriptionb

Emission ratec

g/d lb/d Exit temp. (K); control device

Main plant

Control room

1. Blending, slugging,

compacting, granulating

6.12 0.0135 297; Baghouse

2. Slugging, granulating 1.22 0.0027 297; Baghouse

3. Pelleting, consolidating 1.63d 0.0036d 295; Baghouse

4. Pelleting, consolidating 42.46 0.0936 297; Baghouse

4a. Pelleting, consolidating 6.53 0.0144 297; Baghouse

5. Blending, compacting,

granulating, pelleting,

consolidating

1.36d 0.003b 297; Baghouse

Anode room

6. Amalgam, dewatering 1.82d 0.004d 297; Uncontrolled

6a. Vacuum dryer 0.46d 0.001d 297; Uncontrolled

6b. Blending 0.91d 0.002d 297; Uncontrolled

7. Pelleting, zinc amalgam 4.08d 0.009d 295; Baghouse

Cell assembly area

8. Assembling calls 28.58 0.0630 295; Baghouse for PM. Vapor by

recirculating air through prefilters

and charcoal filters

 U.S. EPA, 1984. a

 Source names are those used by facility. b

 Emission rates were measured by facility except where noted. c

 Estimated emission rate by facility. d
4-68

Table 4-21

1996 U.S. Byproduct Mercury-Producing Gold Minesa

Mine County/State Operator

Alligator Ridge White Pines, NV Placer Dome U.S.

Getchell Humboldt, NV FMC Gold Co.

Carlin Mines Complex Eureka, NV Newmont Gold Co.

McLaughlin Napa, CA Homestake Mining Co.

Mercur Tooele, UT Barrick Mercur Gold Mines, Inc.

Pinson Mine Humboldt, NV Pinson Mining Co.

 Plachy, 1997. a

4.2.13 Primary Mercury Production

Mercury is currently only produced in the United States as a byproduct from the mining of gold

ores and is no longer produced from mercury ore. The last U.S. mercury ore mine, the McDermitt Mine

in McDermitt, Nevada, ceased operation in 1990, and all its equipment has since been dismantled, sold,

landfilled, or scrapped (U.S. EPA, 1997a).

Since the closure of the McDermitt Mine, recovery of mercury as a byproduct from gold ores is

the only remaining ore-based production process. In 1996, six U.S. gold mines (four in Nevada, one in

California and one in Utah) produced metallic mercury as a byproduct. Mines that do produce mercury

represent only a small percentage of all domestic gold mines. The names and locations of these mines

are shown in Table 4-21. No information was available on the amount of mercury recovered at each

facility, although the Bureau of Mines reported that 64 Mg (70 tons) of mercury was produced as a

byproduct of gold ore mining in 1992 (Bureau of Mines, 1994). Data are insufficient at this time to

estimate the quantity of mercury emissions generated as a byproduct of gold ore mining.

Potential sources of mercury emissions from gold processing facilities are at locations where furnaces,

retorts, or other high-temperature sources are used in the process and where the mercury is removed from

the launders. The treated gas discharged to the atmosphere is also a source of mercury emissions (U.S.

EPA, 1997a).
4-69

Primary Mercury Production Processes

This description of production processes and emission controls used at gold mines does not necessarily reflect any

specific gold mine but summarizes the types of processes and controls a gold mine could use to produce mercury and

control mercury emissions. These processes vary from site to site.

The incoming gold ore is crushed using a series of jaw crushers, cone crushers and ball mills. If the incoming ore is

an oxide-based ore, no pretreatment is required and the crushed ore is mixed with water and sent to the classifier. If the

ore is a sulfide-based ore, it must be pretreated using either a fluid bed or multiple hearth pretreatment furnace (roaster) to

convert metallic sulfides to metallic oxides. The exhaust gas from either of these units is sent through wet ESPs and, if

necessary, through carbon condensers. The exhaust gas then passes through a lime sulfur dioxide (SO ) scrubber prior to 2

discharging to the atmosphere. If the treated sulfide ore is high in mercury content, the primary mercury recovery process

occurs from the wet ESPs. If the concentration is low, no attempt is made to recover the mercury for sale. The pretreated

ore is mixed with water and sent to the classifier, where the ore is separated (classified) according to size. Ore pieces too

large to continue in the process are returned to the crusher operation (U.S. EPA, 1993a).

From the classifier, the slurry passes through a concentrator and then to a series of agitators containing the cyanide

leach solution. From the agitators, the slurry is filtered, the filter cake sent to disposal and the filtrate containing the gold

and mercury is transferred to the electrowinning process. If the carbon-in-pulp (CIP) process is used, the cyanide pulp in

the agitators is treated with activated carbon to adsorb the gold and mercury. The carbon is filtered from the agitator

tanks and treated with an alkaline cyanide-alcohol solution to desorb the metals. This liquid is then transferred to the

electrowinning tanks. In the electrowinning process, the gold and mercury are electrodeposited onto a stainless steel wool

cathode, which is sent to a retort to remove mercury and other volatile impurities. The stainless steel wool, containing the

gold, is transferred from the retort to a separate smelting furnace, where the gold is volatilized and recovered as crude

bullion (U.S. EPA, 1993a).

The exhaust gas from the retort, containing mercury, SO , PM, water vapor, and other volatile components, passes 2

through condenser tubes, where the mercury condenses as a liquid and is collected under water in the launders. From the

launders, the mercury is purified and sent to storage. After passing through the condenser tubes, the exhaust gas goes

through a venturi and impinger tower to remove PM and water droplets and then moves through the SO scrubber prior to 2

discharging to the atmosphere (U.S. EPA, 1993a).

When pretreatment roasting is required, the exhaust gases from the furnace pass through a cyclone to remove PM

and then move through wet ESPs to remove arsenic, mercury and some of the SO . If the mercury concentration in the 2

gold ore is high, the ESPs will not remove all of the mercury, and an activated carbon adsorber bed may be required for

additional mercury removal. The gas passes through a lime scrubber to remove SO ; if the SO concentration is low, a 2 2

caustic scrubber may be used. From the scrubber, the gas is discharged through the stack to the atmosphere. Essentially,

the same emission control measures are used for the exhaust gas from the retort. After the gas passes through the

condenser tubes to remove the mercury, a venturi and a cyclone are used to remove PM and water droplets. These

controls are followed by the lime scrubber to remove the SO prior to discharging to the atmosphere. 2

Gold ores in open heaps and dumps can also be treated by cyanide leaching. In this process, the gold ore is placed

on a leaching pad and sprayed with the cyanide solution. The solution migrates down through the ore to a collection

system on the pad and then is sent to a pregnant solution pond. From this pond, the leachate liquors, containing gold and

mercury, are transferred to the gold recovery area. In this area, the liquor is filtered and sent to the electrowinning

process (U.S. EPA, 1993a).

No emission data have been published for facilities producing mercury as a byproduct of gold

ore; therefore, no estimate of mercury emissions from gold ore mining can be made at this time. 

According to an industry representative, all gold mines that produce mercury control their emissions

because the objective is to recover as much mercury as possible (Barringer and Johnson, 1995).
4-70

Table 4-22

1995 U.S. Mercury Compound Producersa

Producer Location

1991 TRI

emissions, 

kg (lb)b Compound(s)

Elf Atochem North America, Inc.,

Chemical Specialties Division

Tulsa, OK Nrc HgF Hg F 2, 2 2

GFS Chemicals, Inc. Columbus, OH NR HgBr , HgI , Hg(NO ) , 2 2 32

HgSO4

Johnson Matthey, Inc. Ward Hill, MA NR Hg (NO ) 2 32

R.S.A Corporation Danbury, CT NR Hg(SCN)2

 SRI, 1996. a

 U.S. EPA, 1996b. b

 NR = Not reported; company did not appear in 1994 TRI. c

No specific data on emission factors from potential sources of mercury emissions from mercury

ore mining have been published since 1973 (U.S. EPA, 1997a). The 1973 report gives a total emission

factor of 0.171 kg of mercury emitted for each megagram of mercury ore mined (0.342 lb/ton), which

was based on stack tests conducted in the early 1970s (Anderson, 1973). However, this emission factor

is for mercury emissions from mercury ore mining only and cannot be used for mercury emissions from

gold ore mining. No mercury emissions from gold ore mining were, therefore, estimated for this report.

4.2.14 Mercury Compounds Production

The production of mercury compounds presents a potential source of mercury emissions into the

atmosphere. Common mercury compounds include mercuric chloride and mercuric oxide. Table 4-22

presents a list of several producers of inorganic and organic mercury compounds.

Because numerous mercury compounds are produced in the United States, it is beyond the scope of this

study to present process descriptions for each one. Process descriptions of the more common mercury

compounds can be found in the mercury L&E document (U.S. EPA, 1997a).

During the production of mercury compounds, emissions of mercury vapor and particulate

mercury compounds may occur at the following sources: reactors, driers, filters, grinders, and transfer

operations. No information was found on specific emission control devices to remove or treat the

mercury emissions, but the literature did contain information on methods designed to reduce the

workplace concentrations without subsequent treatment (Reisdorf and D'Orlando, 1984). Typically,

these procedures included some combination of enclosure or containment, process modifications,

exhaust ventilation, dilution ventilation, and personal protective equipment (Reisdorf and D'Orlando,

1984). In some cases, ventilation systems are reported to be ducted to cyclone dust collectors to reduce

dust emissions, but no information was located on mercury vapor controls (U.S. EPA, 1997a). No

information was available from the literature on mercury emissions or emission factors from the
production of mercury compounds; therefore, no mercury emission estimate could be developed. As

shown in Table 4-22, no company reported significant emissions in the 1994 TRI.

4.2.15 Bvoroduct Coke Production

Byproduct coke, also called metallurgical coke, is a primary feedstock for the integrated iron

and steel industry. Because no information concerning mercury emissions from the production of

byproduct coke could be found in the literature, no nationwide mercury emission estimates were

generated. Table 4-23 lists U.S. byproduct coke oven facilities in 1991 (Huskanen, 1991) and Figure

4-16 shows the locations of these facilities.

Coke is currently produced in two types of coke oven batteries: the slot oven byproduct

battery and the nonrecovery battery. The slot oven byproduct type is by far the most commonly used

battery; over 99 percent of coke produced in 1990 was produced in this type of battery (Easterly et al.;

U.S. EPA, 1988).

The byproduct coke oven battery consists of a series (ranging from 10 to 100) of narrow ovens,

0.4 to 0.6 m (1.3 to 2 ft) wide, and 12 to 18 m (40 to 60 ft) long. The height of the ovens may range

between 3 and 6 m (10 and 20 ft). Depending on the dimensions, the production capacity may range

between 6.8 and 35 Mg (7.5 and 39 tons) of coke per batch. A heating flue is located between each

oven pair (Easterly et al.; U.S. EPA, 1988).

Figure 4-16

1991 U.S. Byproduct Coke Producers

i’ /j i;-” ” Li /

i &_ ,__-.‘,‘----.--__-___~’ -----.---7

‘, ._ \A I

L__------‘, hi w ----! -----___ _ __._ \

\‘.

0 L._-I---- ..__1

-1 ‘i !,‘..

/’

‘r--u.--fw*

i -----q ---.__.-1 ‘,o *’

: ‘\* l,

“,_-4<$~

‘_iLUd p* 1

/ -.-._ ,L_ J"; \~>&..:’

------_ _.... -_-_-----I i;’ p- .___l c______ _? ;- -J--------

4-71
4-72

Table 4-23

1991 U.S. Byproduct Coke Producersa

Facility

No. of

batteries

Total No. of

ovens

Total capacity, Mg/d

(ton/d)

Acme Steel, Chicago, IL 2 100 1,450 (1,600)

Armco, Inc., Ashland, KY 2 146 2,450 (2,700)

Armco, Inc., Middleton, OH 3 203 4,130 (4,540)

Bethlehem Steel, Bethlehem, PA 3 284 3,580 (3,940)

Bethlehem Steel, Burns Harbor, IN 2 164 3,980 (4,380)

Bethlehem Steel, Lackawanna, NY 2 152 1,700 (1,870)

Bethlehem Steel, Sparrows Point, MD 3 210 3,700 (4,070)

Geneva Steel, Orem, UT 1 208 2,050 (2,250)

Gulf States Steel, Gadsden, AL 2 130 2,550 (2,800)

Inland Steel, East Chicago, IN 6 446 5,250 (5,780)

LTV Steel, Pittsburgh, PA 5 315 4,910 (5,400)

LTV Steel, Chicago, IL 1 60 1,450 (1,600)

LTV Steel, Cleveland, OH 2 126 2,910 (3,200)

LTV Steel, Warren, OH 1 85 1,360 (1,500)

National Steel, Granite City, IL 2 90 1,380 (1,520)

National Steel, Ecorse, MI 1 78 840 (925)

USS, Div. of USX Corp., Clairton, PA 12 816 11,490 (12,640)

USS, Div. of USX Corp., Gary, IN 6 422 6,490 (7,140)

Wheeling-Pittsburgh Steel, East Steubenville,

WV

4 224 3,450 (3,800)

Total 58 4,259 65,120 (71,660)

 Huskanen, 1991. a
4-73

Pulverized coal, which is the feedstock, is fed through ports located on the top of each oven by a

car that travels on tracks along the top of each battery. The ports are sealed upon charging, and gaseous

fuel is combusted in the flues located between the ovens to provide the energy for the pyrolysis. The

coking process takes between 12 and 20 hours, at the end of which almost all the volatile matter from the

coal is driven off, thus forming coke. The coke is then unloaded from the ovens through vertical doors

on each end of the oven into a rail car, where it is cooled by being sprayed with several thousand gallons

of water. The rail car then unloads the coke in a separate area, where the coke is allowed to cool further

(Easterly et al.; U.S. EPA, 1988).

Mercury is present in coal in appreciable quantities. Consequently, the volatile gases that evolve

from the coking operation are likely to contain mercury (Easterly et al.; U.S. EPA, 1988).

Emissions at byproduct coke plants are generated during coal preparation, oven charging

operations and other operations. Emissions are also generated from door leaks and from the battery

stack. The battery stack emissions are primarily a result of leakage from the oven into the flue. Mercury

emissions can be generated in small quantities during coal preparation and handling as fugitive PM

because mercury is present as a trace contaminant in coal. Mercury also may be volatilized and released

during charging and pushing operations as well as from the battery stacks and door and topside leaks.

There are no mercury data for coke ovens in the U.S., so an estimate of U.S. mercury emissions

from this source category is not included in this report. There are European emission factors available

however, so a rough estimate can be calculated if only to give a sense of the potential magnitude of this

source category's emissions. Emission factors used in Germany for coke production range from 0.01 to

0.03g mercury per Mg of coke produced (Jockel and Hartje, 1991). One difference between European

coke producers and U.S. coke producers is that U.S. coke producers use a very high quality cleaned coal

while their European counterparts do not. If it is assumed that an emission factor of about 0.025 g

mercury per Mg of coke produced is relevant for the U.S. (assuming a 20 percent reduction of mercury

by the coal cleaning process), then potential mercury emissions for this source category would be 0.6

Mg/year (0.7 tons/year).

4.2.16 Petroleum Refining

Petroleum refining involves converting crude petroleum oil into refined products, including

liquified petroleum gas, gasoline, kerosene, aviation fuel, diesel fuel, fuel oils, lubricating oils, and

feedstocks for the petroleum industry. Mercury is reported to be present in petroleum crude, with its

content ranging from 0.023 to 30 ppmwt (U.S. EPA, 1990).

As of January 1995, there were 34 oil companies in the United States with operable atmospheric

crude oil distillation capacities in excess of 100,000 barrels per calendar day. These oil companies

operated refineries at a total of 107 different locations. In addition, there are 53 companies with

distillation capacities of less than 100,000 barrels per calendar day (National Petroleum Refiners

Association, 1995).

The operations at refineries are classified into five general categories: separation processes,

petroleum conversion processes, petroleum treating processes, feedstock and product handling, and

auxiliary facilities. In the separation process, crude oil is separated into its constituents (including

paraffinic, naphthionic and aromatic hydrocarbon compounds) by either atmospheric distillation, vacuum

distillation, or gas processing (recovery of light ends). Conversion processes include cracking, coking

and visbreaking, which breaks large molecules into smaller molecules; isomerization and reforming
4-74

processes to rearrange the structures of molecules; and polymerization and alkylation to combine small

molecules into larger ones (U.S. EPA, 1997a).

Petroleum treatment processes include hydrodesulfurization, hydrotreating, chemical

sweetening, acid gas removal, and deasphalting. These treatment methods are used to stabilize and

upgrade petroleum products. Feedstock and product handling includes storage, blending, loading, and

unloading of petroleum crude and petroleum products. Auxiliary facilities include boilers, gas turbines,

wastewater treatment facilities, hydrogen plants, cooling towers, and sulfur recovery units (U.S. EPA,

1997a).

Control of VOC emissions from distillation, catalytic cracking, coking, blowdown system,

sweetening, and asphalt blowing is achieved by flares. In some cases, the VOC-laden gas stream is also

used as fuel in process heaters. Cyclones in conjunction with ESPs are used to reduce emissions from

catalytic cracking (U.S. EPA, 1997a). These control measures are expected to have little effect on

mercury emissions.

The primary source of mercury emissions in petroleum refining is the separation process,

although mercury emissions can also be expected in the petroleum conversion and petroleum treating

processes (U.S. EPA, 1997a). Data were unavailable, however, to calculate an emission factor. As a

result, no estimate of mercury emissions could be made for this source category. More analyses of oils

and refinery emissions are needed to evaluate this source.

4.3 Miscellaneous Sources

Sources not readily classified as combustion or manufacturing sources of mercury or that once

emitted mercury but currently do not are considered miscellaneous sources. These sources account for

an estimated 1.3 Mg/yr (1.4 tons/yr) of mercury emissions generated in the United States. They include

geothermal power plants, pigments, oil shale retorting, mercury catalysts and explosives. Table 4-24

presents mercury emissions from these miscellaneous sources.

4.3.1 Geothermal Power Plants

Geothermal power plants are either dry-steam or water-dominated and emitted an estimated 1.3

Mg (1.4 tons) of mercury in 1993. For dry-steam plants, steam is pumped from geothermal reservoirs to

turbines at a temperature of about 180C (360F) and a pressure of 7.9 bars absolute (U.S. EPA, 1997a). 

For water-dominated plants, water exists in the producing strata at a temperature of approximately

270C (520F) and at a pressure slightly higher than hydrostatic (U.S. EPA, 1997a). As the water flows

towards the surface, pressure decreases and steam is formed, which is used to operate the turbines. As of

1992, there were 18 geothermal power plants operating in the United States (Marshal, 1993), and one

new plant began operating in 1993 (IGA, 1995). Table 4-25 lists the names, locations and capacities of

these facilities.

No data on the mercury content of steam or water cycled through geothermal facilities are

available. Likewise, no information exists on emission control systems for geothermal power plants

(U.S. EPA, 1997a).
4-75

Table 4-24

Best Point Estimates of Mercury Emissions from Miscellaneous Anthropogenic Emission Sources: 19

Source Data Mg/yr Tons/yr % of Total Uncertainty Basis for Em

Emissions

Date of Degree of

a b

Geothermal power plants 1.3 1.4 0.9 1977/1992 High Test dataTurf products - - - - - No active registratio

mercury-containing 

Pigment production - - - - - No sources in U.S.Oil shale retorting - - - - - No sources in U.S.Mercury catalysts - - - - - Insufficient informa

emissions

Explosives manufacturing - - - - - No sources in U.S.Total 1.3 1.4 0.9

 Date that data emission factor is based on/Date of activity factor used to estimate emissions. a

 A "medium" degree of uncertainty means the emission estimate is believed to be accurate within + 25 percent. A "high" degree of uncertainty means the emission esti

b

within + 50 percent.
4-76

Table 4-25

1992 U.S. Geothermal Power Plantsa

Facility Type Net Capacity (MW)

The Geysers, CA

Salton Sea, CA

Heber, CA

East Mesa, CA

Coso, CA

Casa Diablo, CA

Amedee, CA

Wendel, CA

Puna, HI

Dixie Valley, NV

Steamboat Hot Springs, NV

Beowawe Hot Springs, NV

Desert Peak, NV

Wabuska Hot Springs, NV

Soda Lake, NV

Stillwater, NV

Empire and San Emidio, NV

Roosevelt Hot Springs, UT

Cove Fort, UT

Dry-steam

Water-dominated

Water-dominated

Water-dominated

Water-dominated

Water-dominated

Water-dominated

Water-dominated

Not specified

Water-dominated

Water-dominated

Water-dominated

Water-dominated

Water-dominated

Water-dominated

Water-dominated

Water-dominated

Water-dominated

Water-dominated

1,805.7

218.3

47.0

106.0

247.5

34.0

2.0

0.7

25.0

57.0

19.3

16.7

9.0

1.7

15.7

12.5

3.2

20.0

12.1

Total 2,653

 Marshal, 1993, for all data except for Puna, Hawaii data. Puna data from International Geothermal Association, 1995. a

Puna facility began operating in 1993.

Mercury emissions at geothermal power plants are documented to result from two sources: offgas ejectors and cooling towers. Table 4-26 contains the mercury emission factors for these two 

sources, which are based on measurements taken in 1977 (Robertson et al., 1977). No process data are

given in the documentation containing the test results, and the primary draft source of these data could

not be obtained in order to verify the validity of the emission factors (U.S. EPA, 1997a). If significant

process modifications or changes in control strategies have been incorporated since 1977, the emission

factors reported in Table 4-26 may no longer be valid.

Multiplying the emission factors in Table 4-26 by the total capacity shown in Table 4-24

(assuming that geothermal power plants operate 24 hr/d, 365 d/yr) gives a mercury emission estimate of

1.3 Mg (1.4 tons) for geothermal power plants in 1993. Because the emission factors used to generate

this estimate have limited reliability, this emission estimate has a high degree of uncertainty.
4-77

Table 4-26

Mercury Emission Factors for Geothermal Power Plants a

Source g/MWe/hr g/MWe/hr lb/MWe/hr

Emission factor range Average emission factor

Off-gas ejectors 0.00075 - 0.02 0.00725 0.00002

Cooling tower exhaust 0.026 - 0.072 0.05 0.0001

 Robertson et al., 1977. a

4.3.2 Pigments, Oil Shale Retorting, Mercury Catalysts, Turf Products and Explosives

Pigments, oil shale retorting, mercury catalysts, turf products and explosives were once sources

of mercury emissions but no longer. Domestic production of mercury-containing pigments ceased in

1988 (U.S. EPA, 1992a). There are currently no oil shale retorts in the United States (U.S. EPA, 1981). 

As of 1994, there are no active registrations of mercury-containing turf products in the United States. 

All registrations have been canceled or are in the process of cancellation following voluntary

cancellation by the registrants. No emissions of mercury from production mercury catalysts could be

accounted for (U.S. EPA, 1997a). Commercial mercury use in explosives ceased prior to 1970 (U.S.

EPA, 1992a).
5-1

5. EMISSIONS SUMMARY

Table 5-1 summarizes the estimated national mercury emission rates by source category. These

emissions estimates should be regarded as best estimates given available data.

The emissions data presented in this document served three primary purposes. First, the

inventory identifies source categories that emit a significant amount of mercury. This information will

be useful for decision makers when selecting potential candidates for mercury emissions reductions and

in evaluating possible control technologies or pollution prevention measures that could be used to

achieve emission reductions. Second, the inventory was used to identify source types with the potential

to have public health or environmental impacts when evaluated as singular point sources. The source

types so identified were modeled in the local impact analysis to assess the potential public health and

environmental impacts from a single source. Third, the emissions data summarized in this document

served as input to U.S. EPA's long-range transport model which assessed the nationwide dispersion and

deposition of mercury from all of the identified mercury sources in the U.S. The local impact analysis

and long-range transport modeling are described in detail in Volume III of the Mercury Study Report to

Congress -- An Assessment of Exposure From Anthropogenic Mercury Emissions in the United States.

Accuracy of the Inventory

The accuracy of the emission estimates is obviously a factor in assessing the inventory's

usefulness for its intended purposes. Considering the admitted gaps in the inventory, the external peer

review panel that reviewed this work in January 1995 concluded that the missing sources could

contribute as much as 20 percent more mercury emissions to the U.S. total. For comparison, one

reviewer submitted data on the amount of mercury emitted per person in some European countries (based

on anthropogenic emissions only).

Based on the inventory presented in this document, the U.S. inventory represents 0.55 g mercury

per person per year. Based on data submitted during the 1995 external peer review process, 0.90 g

mercury per person per year is emitted in the United Kingdom. In Germany (Western area), 0.75 g

mercury per person per year is emitted. In Poland, 0.88 g mercury per person per year is estimated to be

emitted. The European emission average is about 1.2 g mercury per person per year (Pacyna, 1995). 

This national inventory of estimated mercury emissions compares favorably with other national

estimates. Porcella, et al. (1995) estimated 1990 U.S. mercury emissions to be 154.1 Mg and Pai, et al.

(1997) estimated 1990 emissions at 146.4 Mg. This study estimates the 1994-1995 national baseline

emissions to be 145 Mg. In general, each of these studies used similar emissions estimation techniques

and data sources, and estimates for individual source categories are close. Like this study, these other

studies also used “top down” techniques based on emission factors (e.g., lbs mercury emitted per unit of

energy or lbs product produced) multiplied by an activity level (e.g., pounds product produced in a year). 

This approach is common, particularly for a national estimate where adding up actual emissions from

every source would be unrealistic.

A regional inventory being compiled by the Northeast States for Coordinated Air Use

Management (NESCAUM) was used for a regional modeling study of mercury emissions and 
5-2

Table 5-1

Best Point Estimates of 1994-1995 National Mercury Emission Rates by Category

Sources of mercury Mg/yr tons/yr Inventory a

1994-1995 1994-1995 % of Total bb b

Area sources 3.1 3.4 2.2

 Lamp breakage 1.4 1.5 1.0

 General laboratory use 1.0 1.1 0.7

 Dental preparations 0.6 0.7 0.4

 Landfills 0.07 0.08 0.1

 Mobile sources c c c

 Paint use c c c

 Agricultural burning c c c

Point Sources 141.0 154.7 97.8

 Combustion sources 125.3 137.7 86.9

 Utility boilers

 Coal

 Oil

 Natural gas

 MWCsh

 Commercial/industrial boilers

 Coal

 Oil

 MWIsh

 Hazardous waste combustors e

 Residential boilers

 Oil

 Coal

 SSIs

 Wood-fired boilersf

 Crematories

 Manufacturing sources

 Chlor-alkali

 Portland cement e

 Pulp and paper manufacturing

 Instruments manufacturing

 Secondary Hg production

 Electrical apparatus

 Carbon black

 Lime manufacturing

 Primary lead

 Primary copper

 Fluorescent lamp recycling

 Batteries

 Primary Hg production

 Mercury compounds 

 Byproduct coke

 Refineries

 Miscellaneous sources

 Geothermal power

 Turf products

 Pigments, oil, etc.

TOTAL

47.2 51.8 32.8

(47) (51.6) (32.6) d

(0.2) (0.2) (0.1)

(<0.1) (<0.1) (0.0)

26.9 29.6 18.7

25.8 28.4 17.9

(18.8) (20.7) (13.1)

(7.0) (7.7) (4.9)

14.6 16.0 10.1

6.4 7.1 4.4

3.3 3.6 2.3

(2.9) (3.2) (2.0)

(0.4) (0.5) (0.3)

0.9 1.0 0.6

0.2 0.2 0.1

<0.1 <0.1 0.0

14.4 15.6 10.0

6.5 7.1 4.5

4.4 4.8 3.1

1.7 1.9 1.2

0.5 0.5 0.3

0.4 0.4 0.3

0.3 0.3 0.2

0.3 0.3 0.2

0.1 0.1 0.1

0.1 0.1 0.1

<0.1 <0.1 0.0

<0.1 <0.1 0.0

<0.1 <0.1 0.0

cc c

cc c

cc c

cc c

1.3 1.4 0.9

1.3 1.4 0.9

gg g

gg g

144 158 100

 MWC=Municipal waste combustor; MWI=medical waste incinerator; SSI=sewage sludge incinerator a

Numbers do not add exactly because of rounding. b 

 Insufficient information to estimate 1995 emissions. c

Parentheses denote subtotal within larger point source category d 

 For the purposes of this inventory, cement kilns that burn hazardous waste for fuel are counted as hazardous waste combustors. e

 Includes boilers only; does not include residential wood combustion (wood stoves). f

Mercury has been phased out of use. g

 EPA has finalized emissions guidelines for these source categories which will reduce mercury emissions by at least an additional 90 percent h

over 1995 levels. 
5-3

dispersion in Connecticut, Maine, Maryland, New Hampshire, New Jersey, New York, Rhode Island, and

Vermont. Emissions for each state were allocated to modeling grid cells for regional modeling. A

comparison of the emissions inventory for each of these states to this study’s emission inventory for the

same states produced good agreement. The EPA’s emission inventory is about 19 Mg/year for the

NESCAUM states, while the states’ own estimates total about 16 Mg/year. The state estimates are likely

to be more accurate because in many cases, emissions testing is required for air pollution permits and

these test data were available to the states to estimate emissions from specific facilities (compared to the

EPA’s emission factor approach).

Use of the Inventory for the Local Impact and Control Technology Analyses

While the emission estimates have limitations, they do provide insight into the relative

magnitude of emissions from different groups of sources. Table 5-2 shows the distribution of estimated

emissions among the four major classes of sources of anthropogenic emissions (area sources, combustion

point sources, manufacturing point sources, and miscellaneous point sources).

Of the estimated 144 Megagrams (Mg) (158 tons) of mercury emitted annually into the

atmosphere by anthropogenic sources in the United States, approximately 87 percent is from combustion

point sources, 10 percent is from manufacturing point sources, 2 percent is from area sources, and

1 percent is from miscellaneous sources. Four specific source categories account for approximately

80 percent of the total anthropogenic emissions--coal-fired utility boilers (33 percent), municipal waste

combustion (19 percent), commercial/industrial boilers (18 percent), and medical waste incinerators

(10 percent).

Based on this information, four source categories were selected for the local impact analyses in

Volume IV of this report and the control technology assessment described in Volume VIII of this report. 

The source categories were selected based on the magnitude of their mercury emissions either in the

aggregate as a source category or as single point sources. The source categories were coal- and oil-fired

utility boilers, municipal waste combustors, medical waste incinerators, and chlor-alkali plants. Model

plants representing these categories were developed for both the local impact analyses and the control

technology assessment. The model plants for the local impact analyses are described in detail in

Appendix C to Volume III of this report and for the control technology assessment, in Appendix B of

Volume VIII of this report.

Use of the Inventory for the Long-range Transport Analysis

For the long-range transport analysis, the emissions inventory was mapped for the continental

U.S. The continental U.S. was divided into 40-km square grid cells and the magnitude of the mercury

emissions were calculated for each cell. For the most part, the location (at least to the city level) of the

mercury point sources described in this document were known. 

For area sources where the sources are small, diffuse and numerous, exact locations were not

known. There were a number of source categories where this was the case. The emissions for these

source categories were allocated or apportioned to each county in the U.S. based on a variety of

information. The area sources and the method used to allocate their emissions are shown in Table 5-3.
5-4

Table 5-2

Best Point Estimates of Mercury Emissions from Anthropogenic Sources: 1994-1995

Source Mg/yr Tons/yr % 

Emissions

Area 3.1 3.4 Combustion 125.3 137.7 Manufacturing 14.4 15.6 Miscellaneous 1.3 1.4 Total Inventory 144 158 
5-5

Table 5-3

Mercury Area Sources Allocation Methodology

Area Source Category Emissions

Mg/year

(tons/yr)

Allocation Method

Mercury Lamp Breakage 1.4 (1.5) Nationwide estimate allocated to counties

on a per capita basis (1990 Census).

General Laboratory Usage 1.0 (1.1) Nationwide estimate allocated on a per

capita basis (1990 Census data).

Dental Preparation 0.6 (0.7) Nationwide estimate allocated to counties

on a per capita basis (1990 Census).

Residential Coal Combustion 0.4 (0.5) Nationwide estimate allocated by State

based on fuel consumption (U.S.

Department of Energy, 1996). 

Apportionment to counties within State on

a per capita basis.

Residential Oil Combustion 2.9 (3.2) Nationwide estimate allocated by State

based on fuel consumption (U.S.

Department of Energy, 1996). 

Apportionment to counties within State on

a per capita basis.

Industrial/Commercial

Boilers

 Coal

 Oil

18.8 (20.7)

7.0 (7.7)

Nationwide estimates allocated by State

based on fuel consumption (U.S.

Department of Energy, 1996). 

Apportionment to counties within State on

a per capita basis.

Crematories < 0.1 Nationwide emissions estimate allocated

to counties on a per capita basis.

Figure 5-1 illustrates the spatial distribution of mercury emissions across the U.S. based on this

inventory. This distribution formed the basis of the long-range transport modeling and the resulting

predictions of wet and dry deposition across the U.S.

5-7

Trends in Mercury Emissions

It is difficult to predict with confidence the temporal trends in mercury emissions for the U.S.,

although there appears to be a trend toward decreasing total mercury emissions from 1990 to 1995. This

is particularly true for the combustion sources where emissions have declined 50 percent from municipal

waste combustors and 75 percent from medical waste incinerators since 1990 (see below). Also, as

previously noted, there are a number of source categories where there is insufficient data to estimate

current emissions let alone potential future emissions. Based on available information, however, a

number of observations can be made regarding mercury emission trends from source categories where

some information is available about past activities and projected future activities.

Current emissions of mercury from manufacturing sources are generally low compared to

combustion sources (with the exception of chlor-alkali plants using the mercury cell process and portland

cement manufacturing plants). The emissions of mercury are more likely to occur when the product (e.g.,

lamps, thermostats) is broken or discarded. Therefore, in terms of emission trends, one would expect that

if the future consumption of mercury remains consistent with the 1996 consumption rate, emissions from

most manufacturing sources would remain about the same.

For industrial or manufacturing sources that use mercury in products or processes, the overall

consumption of mercury is generally declining. Industrial consumption of mercury has declined by

about 75 percent between 1988 (1503 Mg) and 1996 (372 Mg). Much of this decline can be attributed to

the elimination of mercury as a paint additive (20 percent) and the reduction of mercury in batteries (36

percent). Use of mercury by other source categories remained about the same between 1988 and 1996.

Secondary production of mercury (i.e., recovering mercury from waste products) has increased

significantly over the past few years. While 372 Mg of mercury were used in industrial processes in

1996, 446 Mg were produced by secondary mercury producers and an additional 340 Mg were imported. 

This is a two-fold increase since 1991. The number of secondary mercury producers is expected to

increase as more facilities open to recover mercury from fluorescent lamps and other mercury-containing

products (e.g., thermostats). As a result there is potential for mercury emissions from this source

category to increase.

The largest identified source of mercury emissions during 1994-1995 is fossil fuel combustion by

utility boilers, particularly coal combustion. Future trends in mercury emissions from this source

category are largely dependent on both the nation's future energy needs and the fuel chosen to meet those

needs. Another factor is the nature of actions the utility industry may take in the future to meet other air

quality requirements under the Clean Air Act (e.g., national ambient air quality standards for ozone and

particulate matter).

Two other significant sources of mercury emissions currently are municipal waste combustors

and medical waste incinerators. Emissions from these source categories have declined considerably

since 1990 on account of plant closures (for medical waste incinerators) and reduction in the mercury

content of the waste stream (municipal waste combustors) and will decline even further by the year 2000

due to regulatory action the U.S. EPA is taking under the statutory authority of section 129 of the CAA. 

As described in sections 4.1.2 and 4.1.4 of this document, the U.S. EPA has finalized rules for municipal

waste combustors and medical waste incinerators that will, when fully implemented, reduce mercury

emissions from both of these source categories by an additional 90 percent over 1995 levels. In addition

to this federal action, a number of states (including Minnesota, Florida and New Jersey) have

implemented mandatory recycling programs to reduce mercury-containing waste, and some states have

regulations that impose emission limits that are lower than the federal regulation. These factors will

reduce national mercury emissions from these source categories even further.
6-1

6. CONCLUSIONS

The following conclusions are presented in approximate order of degree of certainty in the conclusion,

based on the quality of the underlying database. The conclusions progress from those with greater

certainty to those with lesser certainty.

 Numerous industrial and manufacturing processes emit mercury to the atmosphere. Mercury

emissions from U.S. manufacturing sources, however, have dropped about 75 percent over the

past decade.

 Mercury is emitted, to a varying degree, from anthropogenic sources virtually everywhere in the

United States.

 Natural sources of mercury and re-emission of previously deposited mercury are also sources of

mercury to the atmosphere, although the magnitude of the contribution of these sources relative

to the contribution of current anthropogenic sources is not well understood.

 Prior to 1995, municipal waste combustors and medical waste incinerators were the largest

identifiable source of mercury emissions to the atmosphere. Regulations which have been

finalized for municipal waste combustors and medical waste incinerators will, when fully

implemented, reduce emissions from these source categories by an additional 90 percent over

1995 levels.

 Present emissions estimates indicate that coal-fired utility boilers are the single largest emissions

source, contributing approximately 33 percent of the national inventory.

 Anthropogenic sources in the United States emit approximately 144 Mg (158 tons) of mercury

annually into the atmosphere. This estimate is believed to be accurate to within 30 percent. This

estimate represents emissions calculated during the 1994-1995 time frame. 

 In the United States, areas east of the Rocky Mountains have the highest concentration of

emissions from anthropogenic sources in the U.S.

 The areas having the greatest concentration of mercury emissions from anthropogenic sources of

total mercury (i.e., all chemical species) are the following: the urban corridor from Washington

D.C. to Boston, the Tampa and Miami areas of Florida, the larger urban areas of the Midwest

and Ohio Valley and two sites in northeastern Texas.

 The areas having generally the lowest emissions are in the Great Basin region of the western

United States and the High Plains region of the central United States. There are generally few

large emission sources in the western third of the United States, with the exception of the San

Francisco and Los Angeles areas and specific industrial operations.

There are many uncertainties in the emission estimates for individual source categories due to

uncertainties inherent in an emission factor approach. The source of these uncertainties include the

following: 

 Variability in the estimates of source activity for each source category. Activity levels used in

this Report were compiled over different time periods and by a variety of survey procedures.

 Emissions test data that are of poor quality or are based on very few analyses, which may not be

representative of the full source population being studied.
6-2

 Changes in processes or emission measurement techniques over time (especially since about

1985). Earlier techniques may have measured too much mercury because of contamination

problems.

 A lack of data for some source categories which either led to estimates based on engineering

judgment or mass balance calculations. For a number of source categories there were

insufficient data and, thus, no emissions estimates were made.

 Limited data on the effectiveness of air pollution control equipment to capture mercury

emissions.

Understanding the public health and environmental impacts of current anthropogenic emissions is

complicated by an incomplete understanding of the following factors:

 Global and transboundary deposition of mercury and the impact this has on deposition of

mercury in the U.S.

 The magnitude and chemical nature of natural emissions.

 The magnitude and chemical nature of re-emitted mercury.

 The public health and environmental impacts of emissions from past uses of mercury (such as

paint application) relative to current anthropogenic emissions. 

To improve the emissions estimates, U.S. EPA would need the following:

 Source test data from a number of source categories that have been identified in this volume as

having insufficient data to estimate emissions. Notable among these are mobile sources,

agricultural burning, sludge application, coke ovens, petroleum refining, residential woodstoves,

mercury compounds production and zinc mining.

 Improvements in the existing emissions information for a number of source categories including

secondary mercury production (i.e., recycling), commercial and industrial boilers, landfills,

electric lamp breakage, and iron and steel manufacturing.

 Validation of a stack test protocol for speciated mercury emissions.

 More data on the efficacy of conventional coal cleaning and the potential for slurries from the

cleaning process to be a mercury emission source.

 More data are needed on the mercury content of various coals and petroleum and the trends in

the mercury content of coal burned at utilities and petroleum refined in the U.S.

 Additional research to address the potential for methylmercury to be emitted (or formed) in the

flue gas of combustion sources.

 Investigation of the importance (quantitatively) of re-emission of mercury from previously

deposited anthropogenic emissions and mercury-bearing mining waste. This would include both

terrestrial and water environments. Measuring the flux of mercury from various environments

would allow a determination to be made of the relative importance of re-emitted mercury to the

overall emissions of current anthropogenic sources.
6-3

 Determination of the mercury flux from natural sources to help determine the impact of U.S.

anthropogenic sources on the global mercury cycle as well as the impact of all mercury

emissions in the United States.

 More detailed emissions data to support the use of more sophisticated fate and transport models

for mercury; in particular, more information is needed on the chemical species of mercury being

emitted (including whether these species are particle-bound) and the temporal variability of the

emissions.

Based on trends in mercury use and emissions, the U.S. EPA predicts the following:

 A significant decrease (at least 90 percent over 1995 levels) will occur in mercury emissions

from municipal waste combustors and medical waste incinerators when the regulations finalized

by U.S. EPA for these source categories are fully implemented.

 Manufacturing use of mercury will continue to decline with chlorine production from mercury

cell chlor-alkali plants continuing to account for most of the mercury use in the manufacturing

sector.

 Secondary production of mercury will continue to increase as more recycling facilities

commence operation to recover mercury from discarded products and wastes.
7-1

7. RESEARCH NEEDS

Throughout this volume an effort has been made to characterize the uncertainties (at least

qualitatively) in the emissions estimates for the various source categories described. As noted in Chapter

1, there are inherent uncertainties in estimating emissions using emission factors. To reduce these

uncertainties, a number of research needs remain, including the following:

 Source test data are needed from a number of source categories that have been identified

in this volume as having insufficient data to estimate emissions. These source categories

are listed in Table 1-3. Notable among these are mobile sources, agricultural burning,

sludge application, coke ovens, petroleum refining, residential woodstoves, mercury

compounds production and zinc mining. A number of manufacturing sources were also

identified as having highly uncertain emissions estimates. Notable among this category

are secondary mercury production, commercial and industrial boilers, electric lamp

breakage, landfills, primary metal smelting operations and iron and steel manufacturing. 

The possibility of using emissions data from other countries could be further

investigated.

 Development and validation of a stack test protocol for speciated mercury emissions is

needed.

 More data are needed on the efficacy of conventional coal cleaning and the potential for

slurries from the cleaning process to be a mercury emission source.

 More data are needed on the mercury content of various coals and petroleum and the

trends in the mercury content of coal burned at utilities and petroleum refined in the U.S.

 Additional research is needed to address the potential for methylmercury to be emitted

(or formed) in the flue gas of combustion sources.

 The importance (quantitatively) of re-emission of mercury from previously deposited

anthropogenic emissions and mercury-bearing mining waste needs to be investigated. 

This would include both terrestrial and water environments. Measuring the flux of

mercury from various environments would allow a determination to be made of the

relative importance of re-emitted mercury to the overall emissions of current

anthropogenic sources.

 Determination of the mercury flux from natural sources would help determine the impact

of U.S. anthropogenic sources on the global mercury cycle as well as the impact of all

mercury emissions in the United States.

 The use of more sophisticated fate and transport models for mercury will require more

detailed emissions data, particularly more information on the chemical species of

mercury being emitted (including whether these species are particle-bound) and the

temporal variability of the emissions.
8-1

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APPENDIX A

INFORMATION ON LOCATIONS OF AND EMISSIONS FROM

COMBUSTION SOURCES
A-1

Table A-1

1994 Estimated Mercury Emissions From Utility Boilers, By State and Fuel Type

State No. Mg/yr tons/yr No. Mg/yr tons/yr No. Mg/yr tons/yr No. Mg/yr tons/yr 

Combined cycle Coal Natural Gas Oil 

a

Alaska 0 0 0 1 0.005 0.006 0 0 0 0 0 0 Alabama 0 0 0 38 1.892 2.086 2 0.000 0.000 0 0 0 Arkansas 0 0 0 5 0.411 0.453 6 0.000 0.000 0 0 0 Arizona 0 0 0 14 0.706 0.778 14 0.000 0.000 0 0 0 California 4 0.001 0.001 0 0 0 96 0.000 0.000 38 0.008 0.009 Colorado 0 0 0 24 0.830 0.915 3 0.000 0.000 0 0 0 Connecticut 0 0 0 1 0.071 0.079 2 0.000 0.000 19 0.010 0.011 Delaware 0 0 0 6 0.137 0.151 2 0.000 0.000 2 0.001 0.001 Florida 2 0.001 0.001 29 1.258 1.387 51 0.000 0.000 49 0.062 0.069 Georgia 0 0 0 36 1.753 1.932 6 0.000 0.000 2 0.000 0.000 GU 0 0 0 0 0 0 0 0 0 4 0.002 0.003 Hawaii 0 0 0 0 0 0 0 0 0 14 0.017 0.010 Iowa 0 0 0 31 0.780 0.860 0 0 0 0 0 0 Illinois 0 0 0 57 1.265 1.395 5 0.000 0.000 10 0.002 0.002 Indiana 0 0 0 71 2.195 2.420 2 0.000 0.000 5 0.000 0.000 Kansas 0 0 0 19 0.466 0.514 17 0.000 0.000 0 0 0 Kentucky 0 0 0 55 1.766 1.946 0 0.000 0.000 1 0.000 0.000 Louisiana 2 0.000 0.000 6 0.737 0.812 46 0.000 0.000 0 0.000 0.000 Massachusetts 0 0 0 9 0.288 0.318 8 0.000 0.000 19 0.011 0.013 Maryland 0 0 0 14 0.937 1.033 4 0.000 0.000 13 0.008 0.009 Maine 0 0 0 0 0 0 0 0 0 8 0.003 0.003 
Table A-1 (continued)

1994 Estimated Mercury Emissions From Utility Boilers, By State and Fuel Type

State No. Mg/yr tons/yr No. Mg/yr tons/yr No. Mg/yr tons/yr No. Mg/yr tons/yr 

Combined cycle Coal Natural Gas Oil 

a

A-2

Michigan 0 0 0 52 1.726 1.902 1 0.000 0.000 6 0.002 0.002 

Minnesota 0 0 0 26 0.672 0.741 1 0.000 0.000 0 0.000 0.000 Missouri 0 0 0 34 1.378 1.519 4 0.000 0.000 0 0 0 Mississippi 0 0 0 6 0.173 0.190 15 0.000 0.000 2 0.001 0.001 Montana 0 0 0 6 0.357 0.393 0 0.000 0.000 0 0 0 North Carolina 0 0 0 36 1.253 1.381 0 0 0 0 0 0 North Dakota 0 0 0 13 1.106 1.219 0 0 0 0 0 0 Nebraska 0 0 0 13 0.443 0.488 3 0.000 0.000 0 0 0 New Hampshire 0 0 0 4 0.130 0.144 0 0 0 3 0.002 0.002 New Jersey 0 0 0 8 0.173 0.190 13 0.000 0.000 15 0.005 0.005 New Mexico 0 0 0 10 0.396 0.437 9 0.000 0.000 0 0 0 Nevada 0 0 0 8 0.253 0.278 9 0.000 0.000 6 0 0 New York 0 0 0 32 1.208 1.332 32 0.000 0.000 41 0.057 0.063 Ohio 0 0 0 91 3.613 3.982 1 0.000 0.000 3 0.000 0.000 Oklahoma 4 0.001 0.001 10 0.533 0.587 26 0.000 0.000 0 0 0 Oregon 0 0 0 1 0.034 0.038 0 0 0 0 0 0 Pennsylvania 0 0 0 58 4.657 5.133 2 0 0 8 0.006 0.007 Puerto Rico 0 0 0 0 0 0 0 0 0 18 0.028 0.031 Rhode Island 0 0 0 0 0 0 4 0.000 0.000 1 0.000 0.000 South Carolina 0 0 0 24 0.547 0.603 2 0.000 0.000 4 0.000 0.000 South Dakota 0 0 0 2 0.154 0.170 1 0.000 0.000 0 0 0 
Table A-1 (continued)

1994 Estimated Mercury Emissions From Utility Boilers, By State and Fuel Type

State No. Mg/yr tons/yr No. Mg/yr tons/yr No. Mg/yr tons/yr No. Mg/yr tons/yr 

Combined cycle Coal Natural Gas Oil 

a

A-3

Tennessee 0 0 0 33 1.362 1.501 0 0 0 0 0 0 

Texas 3 0.001 0.001 34 5.599 6.172 154 0.000 0.000 5 0.000 0.000 Utah 0 0 0 12 0.170 0.188 1 0.000 0.000 0 0 0 Virginia 0 0 0 24 0.564 0.621 1 0 0 4 0.002 0.003 VI 0 0 0 0 0 0 0 0 0 2 0.000 0.000 Washington 0 0 0 2 0.181 0.199 0 0 0 0 0.000 0.000 Washington 0 0 0 0 0 0 0 0 0 2 0.001 0.001 D.C.Wisconsin 0 0 0 40 1.063 1.172 2 0.000 0.000 0 0 0 West Virginia 0 0 0 33 1.911 2.107 0 0 0 0 0 0 Wyoming 0 0 0 15 0.851 0.938 0 0 0 0 0 0 Total 15 0.004 0.004 1043 45.999 50.710 545 0.001 0.002 305 0.230 0.250 

 These units burn a combination of fuels. a

Note: Totals shown here differ slightly from those shown elsewhere in this volume due to rounding. For each souce category, a value emissions estimate is zero whereas "0.000" means that the estimate is less than 0.001, the minimum value that could be shown in the s
A-4

Table A-2

Estimates of 1994 Coal, Natural Gas, and Oil Consumption

 in the Commercial/Industrial Sector Per State (Trillion Btu)

State Bituminous and lignite Anthracite Total Natural gas Distillate and residual

Coal

Petroleum

Alabama 146.3 0.1 146.4 227 42.8

Alaska 5.4 a 5.4 356.6 17.9

Arizona 14.6 a 14.6 56.7 12.2

Arkansas 8.6 a 8.6 169.7 25.2

California 56.7 a 56.7 1008.8 69.9

Colorado 18.3 0.5 18.8 162.1 25.3

Connecticut 0.7 0.1 0.8 71.9 31.2

Delaware 4.9 0.3 5.2 23.5 15.9

Dist. of Col. 0.7 a 0.7 14.9 6.4

Florida 32.6 0.2 32.8 188.4 64.8

Georgia 48.9 a 48.9 234.8 44.3

Hawaii 1.8 0.0 1.8 2.3 13.8

Idaho 9.3 0.1 9.4 41.4 17.4

Illinois 152.9 0.2 153.1 513.3 61.8

Indiana 231.0 a 231.0 350.4 56.7

Iowa 54.9 0.7 55.6 157.9 42.3

Kansas 3.8 0.0 3.8 284.6 35.4

Kentucky 87.1 0.3 87.4 130.2 45.1

Louisiana 11.3 0.1 11.4 1278.1 85.3

Maine 11.4 0.1 11.5 4.2 84.3

Maryland 19.3 0.1 19.4 94.6 37.0

Massachusetts 1.7 0.2 1.9 181.7 74.9

Michigan 111.8 a 111.8 538.1 42.0

Minnesota 29.7 0.1 29.8 180.4 49.5

Mississippi 7.1 a 7.1 123.5 27.7

Missouri 27.8 0.1 27.9 138.6 30.3

Montana 10.5 a 10.5 29.9 16.7

Nebraska 8.0 a 8.0 74.9 37.7

Nevada 4.5 0.0 4.5 49.3 18.7

New Hampshire 0.0 0.1 0.1 11.0 20.7

New Jersey 1.6 0.3 1.9 335.5 71.1

New Mexico 1.6 a 1.6 98.5 9.5

New York 75.3 2.2 77.5 450.5 225.5

North Carolina 64.4 a 64.4 138.6 68.4

North Dakota 95.2 a 95.2 29.5 22.1

Ohio 183.4 0.4 183.8 497.0 62.6

Oklahoma 16.1 a 16.1 331.8 25.3

Oregon 2.7 0.2 2.9 89.6 18.5

Pennsylvania 389.5 16.0 405.5 392.8 104.9

Rhode Island 0.0 a a 54.5 13.4

South Carolina 59.4 0.1 59.5 118.9 27.5

South Dakota 8.0 a 8.0 16.4 18.6

Tennessee 103.7 0.4 104.1 175.1 30.6

Texas 82.8 0.1 82.9 2406.3 14.0

Utah 47.7 a 47.7 81.6 15.4

Vermont 0.1 0.0 0.1 4.7 8.3

Virginia 101.9 0.3 102.2 145.2 45.5

Washington 5.1 0.0 5.1 156.7 26.9

West Virginia 112.5 0.1 112.6 84.9 22.8

Wisconsin 48.7 0.1 48.8 216.3 45.0

Wyoming 43.5 a 43.5 93.3 17.7

United States 2564.8 24.6 2589.4 12616.5 2178.1

 Number less than 0.05 a

Source: U.S. Department of Energy. State Energy Data Report. Report No. DOE/EIA-0214(94). October 1996.
A-5

Table A-3

Estimates of Mercury Emissions From Coal-Fired

Commercial/Industrial Boilers on a Per-State Basis For 1994

State

Mercury emissionsa

Ton/Yr Mg/Yr

Alabama 1.2 1.1

Alaska b b

Arizona 0.1 0.1

Arkansas 0.1 0.1

California 0.4 0.4 

Colorado 0.1 0.1 

Connecticut b b

Delaware b b

Dist. of Col. 0.0 b

Florida 0.2 0.2 

Georgia 0.4 0.4 

Hawaii b b

Idaho 0.1 0.1 

Illinois 1.2 1.1 

Indiana 1.9 1.7 

Iowa 0.4 0.4 

Kansas b b

Kentucky 0.7 0.6 

Louisiana 0.1 0.1 

Maine 0.1 0.1 

Maryland 0.1 0.1 

Massachusetts b b

Michigan 0.9 0.8 

Minnesota 0.2 0.2 

Mississippi 0.1 0.1 

Missouri 0.2 0.2 

Montana 0.1 0.1 

Nebraska 0.1 0.1 

Nevada b b

New Hampshire b b

New Jersey b b

New Mexico b b

New York 0.7 0.6 

North Carolina 0.6 0.5 

North Dakota 0.8 0.7 

Ohio 1.4 1.3 

Oklahoma 0.1 0.1 

Oregon b b

Pennsylvania 3.3 3.0 

Rhode Island b b

South Carolina 0.4 0.4 

South Dakota 0.1 0.1 

Tennessee 0.9 0.8 

Texas 0.7 0.6 

Utah 0.3 0.3 

Vermont b b

Virginia 0.8 0.7 

Washington b b

West Virginia 0.9 0.8 

Wisconsin 0.4 0.4 

Wyoming 0.3 0.3 

United States 20.7 18.8 

 Mercury emission factors of 16 lb Hg/trillion Btu and 18 lb Hg/trillion Btu were used for bituminous and anthracite coal, respectively. No a

control of emissions from commercial/industrial boilers was assumed.

 Number less than 0.05. b
A-6

Table A-4

Estimates of Mercury Emissions From Oil-Fired

Commercial/Industrial Boilers On a Per-State Basis For 1994

State Ton/Yr Mg/Yr

Mercury emissionsa

Alabama 0.15 0.14 

Alaska 0.07 0.06 

Arizona 0.04 0.04 

Arkansas 0.09 0.08 

California 0.25 0.23 

Colorado 0.09 0.08 

Connecticut 0.11 0.10 

Delaware 0.06 0.05 

Dist. of Col. 0.02 0.02 

Florida 0.23 0.21 

Georgia 0.15 0.14 

Hawaii 0.04 0.04 

Idaho 0.07 0.06 

Illinois 0.22 0.20 

Indiana 0.20 0.18 

Iowa 0.15 0.14 

Kansas 0.13 0.12 

Kentucky 0.17 0.15 

Louisiana 0.31 0.28 

Maine 0.06 0.26 

Maryland 0.13 0.12 

Massachusetts 0.26 0.24 

Michigan 0.15 0.14 

Minnesota 0.18 0.16 

Mississippi 0.10 0.09 

Missouri 0.11 0.10 

Montana 0.06 0.05 

Nebraska 0.13 0.12

Nevada 0.07 0.06 

New Hampshire 0.08 0.07 

New Jersey 0.25 0.23 

New Mexico 0.03 0.03 

New York 0.78 0.71 

North Carolina 0.24 0.22 

North Dakota 0.08 0.07 

Ohio 0.22 0.20 

Oklahoma 0.09 0.08 

Oregon 0.07 0.06 

Pennsylvania 0.37 0.34 

Rhode Island 0.04 0.04 

South Carolina 0.10 0.09 

South Dakota 0.07 0.06 

Tennessee 0.11 0.10 

Texas 0.50 0.45 

Utah 0.06 0.05 

Vermont 0.03 0.03 

Virginia 0.17 0.15 

Washington 0.10 0.09 

West Virginia 0.08 0.07 

Wisconsin 0.17 0.15 

Wyoming 0.07 0.06 

United States 7.70 7.00 

 Mercury emission factor for distillate oil is 7.2 lb Hg/trillion Btu. Calculation was performed assuming that all pollution control devices a

provide no mercury reduction.
A-7

Table A-5

Estimates of 1994 Coal, Natural Gas, and Oil Consumption

in the Residential Sector Per State (Trillion Btu)

State lignite Anthracite Total Natural gas and residual

Coal

Bituminous coal and Petroleum distillate

Alabama 0.1 a 0.1 51.3 0.1 

Alaska 2.9 0.0 2.9 14.9 7.3 

Arizona a a a 30.5 a

Arkansas a a a 42.4 a

California 1.4 a 1.4 531.7 0.9 

Colorado 0.2 0.0 0.2 99.9 0.1 

Connecticut a 0.2 0.2 42.9 73.2 

Delaware 0.2 a 0.3 8.9 6.9 

Dist. of Col. 0.4 a 0.4 16.0 0.8 

Florida 0.2 a 0.2 15.6 1.5 

Georgia 0.2 a 0.3 108.6 0.7 

Hawaii 0.0 0.0 0.0 0.6 a

Idaho 0.3 a 0.3 12.8 3.1 

Illinois 2.0 a 2.0 483.7 4.7 

Indiana 2.8 0.1 2.8 159.5 10.6 

Iowa 0.3 0.1 0.4 78.9 5.7 

Kansas 0.3 0.0 0.3 74.1 0.2 

Kentucky 2.5 a 2.5 66.4 4.8 

Louisiana 0.0 0.0 0.0 55.0 0.1 

Maine 0.0 0.1 0.1 0.9 32.9 

Maryland 0.3 0.1 0.3 79.0 29.0 

Massachusetts a 0.3 0.3 122.6 115.1 

Michigan 2.5 a 2.5 376.8 23.5 

Minnesota 1.6 a 1.6 123.6 19.7 

Mississippi 0.0 0.0 0.0 27.9 a

Missouri 1.8 a 1.8 123.3 2.1 

Montana a 0.0 a 19.2 1.1 

Nebraska 0.1 0.0 0.1 43.7 0.9 

Nevada a 0.0 a 22.0 0.9 

New Hampshire 0.0 0.1 0.1 6.7 22.2 

New Jersey 0.0 0.2 0.2 225.4 71.9 

New Mexico 0.1 a 0.1 30.9 a

New York 0.8 1.4 2.2 395.9 155.9 

North Carolina 2.3 a 2.3 49.2 19.0 

North Dakota 0.7 0.0 0.7 11.3 4.3 

Ohio 4.2 0.1 4.3 356.0 28.5 

Oklahoma a 0.0 a 71.0 a

Oregon a a a 30.2 5.4 

Pennsylvania 2.2 13.6 15.8 278.1 115.3 

Rhode Island 0.0 a a 17.9 20.5 

South Carolina 0.5 0.1 0.6 24.2 3.9 

South Dakota 0.1 a 0.1 12.2 3.1 

Tennessee 0.8 a 0.8 59.2 1.8 

Texas a a a 222.5 a

Utah 0.9 a 0.9 52.3 0.7 

Vermont a 0.0 a 2.4 12.6 

Virginia 2.7 a 2.8 67.7 28.6 

Washington 0.7 0.0 0.7 55.3 8.9 

West Virginia 0.8 a 0.8 37.5 3.4 

Wisconsin 0.4 a 0.5 129.7 28.0 

Wyoming 1.6 0.0 1.6 12.2 0.4 

United States 38.6 16.4 55.1 4,980.4 880.0 

 Number less than 0.05. a

Source: U.S. Department of Energy. State Energy Data Report. Report No. DOE/EIA-0214(94). October 1996.
A-8

Table A-6

Estimates of Mercury Emissions From

Coal-Fired Residential Boilers on a Per-State Basis For 1994

State Ton/Yr Mg/Yr

Mercury emissionsa

Alabama 0.001 0.001

Alaska 0.023 0.021 

Arizona b b 

Arkansas b b 

California 0.011 0.010 

Colorado 0.001 0.001 

Connecticut 0.001 0.001 

Delaware 0.002 0.002 

Dist. of Col. 0.003 0.003 

Florida 0.001 0.001 

Georgia 0.002 0.002 

Hawaii 0.000 0.000 

Idaho b b 

Illinois 0.017 0.015 

Indiana 0.023 0.021 

Iowa 0.003 0.003 

Kansas 0.002 0.002 

Kentucky 0.020 0.018 

Louisiana 0.000 0.000 

Maine 0.001 0.001 

Maryland 0.003 0.003 

Massachusetts 0.003 0.003 

Michigan 0.020 0.018 

Minnesota 0.012 0.011 

Mississippi 0.000 0.000 

Missouri 0.014 0.013 

Montana b b 

Nebraska b b 

Nevada b b 

New Hampshire 0.001 0.001 

New Jersey 0.001 0.001 

New Mexico b b 

New York 0.019 0.017 

North Carolina 0.019 0.017 

North Dakota 0.006 0.005 

Ohio 0.034 0.031 

Oklahoma b b 

Oregon b b 

Pennsylvania 0.140 0.127 

Rhode Island b b 

South Carolina 0.001 0.001 

South Dakota 0.001 0.001 

Tennessee 0.007 0.006 

Texas b b 

Utah 0.007 0.006 

Vermont b b 

Virginia b b 

Washington 0.022 0.020 

West Virginia 0.006 0.005 

Wisconsin 0.003 0.003 

Wyoming 0.012 0.011 

United States 0.457 0.415 

 Mercury emission factors of 16 lb Hg/trillion Btu and 18 lb Hg/trillion Btu were used for bituminous and anthracite coal, respectively. No a

control of emissions from residential boilers was assumed.

 Number less than 0.05. b
A-9

Table A-7

Estimates of Mercury Emissions From Oil-Fired

Residential Boilers on a Per-State Basis For 1994

State

Mercury emissionsa

Ton/Yr Mg/Yr

Alabama 0.0002 0.0002 

Alaska 0.0263 0.0239 

Arizona 0.0001 0.0001 

Arkansas b b 

California 0.0031 0.0028 

Colorado 0.0006 0.0005 

Connecticut 0.2629 0.2390 

Delaware 0.0246 0.0224 

Dist. of Col. 0.0028 0.0025 

Florida 0.0052 0.0047 

Georgia 0.0023 0.0021 

Hawaii b b 

Idaho 0.0110 0.0100 

Illinois 0.0169 0.0154 

Indiana 0.0383 0.0348 

Iowa 0.0204 0.0185 

Kansas 0.0006 0.0005 

Kentucky 0.0171 0.0155 

Louisiana 0.0002 0.0002 

Maine 0.1180 0.1073 

Maryland 0.1043 0.0948 

Massachusetts 0.4136 0.3760 

Michigan 0.0928 0.0767 

Minnesota 0.0708 0.0644 

Mississippi b b 

Missouri 0.0074 0.0067

Montana 0.0040 0.0036 

Nebraska 0.0034 0.0031 

Nevada 0.0032 0.0029 

New Hampshire 0.0799 0.0726 

New Jersey 0.2583 0.2348 

New Mexico 0.0002 0.0002 

New York 0.5600 0.5091 

North Carolina 0.0750 0.0620 

North Dakota 0.0153 0.0139 

Ohio 0.1024 0.0931 

Oklahoma b b 

Oregon 0.0196 0.0178 

Pennsylvania 0.4143 0.3766 

Rhode Island 0.0736 0.0669 

South Carolina 0.0140 0.0127 

South Dakota 0.0112 0.0102 

Tennessee 0.0064 0.0058 

Texas 0.0001 0.0001 

Utah 0.0023 0.0021 

Vermont 0.0453 0.0412 

Virginia 0.1029 0.0935 

Washington 0.0319 0.0290 

West Virginia 0.0122 0.0111 

Wisconsin 0.1003 0.0912 

Wyoming 0.0014 0.0013 

United States 3.1613 2.8739

 Mercury emission factor for distillate oil is 7.2 lb Hg/trillion Btu. Calculations performed under the assumption that air pollution control a

devices provide no mercury reduction.

 Number less than 0.05. b
A-10

Table A-8

Existing MWC Facilities

Facility County State

Juneau Juneau Burough AK

Sitka (Sheldon Jackson College) Sitka Borough AK

Huntsville Madison/Limestone AL

Batesville Independence AR

Blytheville Incinerator Mississippi AR

Stuttgart Incinerator Arkansas AR 

Osceola Mississippi AR

Commerce Refuse-to-Energy Fac. Los Angeles CA

Long Beach (SERRF) Los Angeles CA

Stanislaus (Modesto) Stanislaus CA

Bridgeport RESCO Fairfield CT

Bristol RRF Hartford CT

MID-Connecticut Project Hartford CT

Town of New Canaan Volume Fairfield CT

Reduction Plane 

Southeastern Connecticut RRF New London CT

Bay Resource Mgt. Center Bay FL

Hillsborough County RRF Hillsborough FL

Broward County RRF North Broward FL 

Broward County RRF South Broward FL

Pasco County Solid Waste RRF Pasco FL

Mayport NAS Duval FL

Dade County Dade FL

Miami International Airport Dade FL

Lake County RR Lake FL

McKay Bay REF Hillsborough FL

Southernmost WTE Monroe FL

Wheelabrator Pinellas RRF Pinellas FL

North Co. Region RR Project West Palm Beach FL

Savannah RRF Chatham GA

Honolulu Resource Recovery Honolulu HI
Table A-8 (continued)

Existing MWC Facilities

Facility County State

A-11

Burley Cassia ID

Northwest WTE Cook & DuPage IL

Indianapolis RRF Marion IN

Springfield RRF Hampden MA

Fall River Incinerator Bristol MA

Haverhill RRF Essex MA

Haverhill Lawrence RDF Essex MA

North Andover RESCO Essex MA

Pittsfield RRF Pittsfield MA

SEMASS RRF Plymouth MA

Saugus RESCO Essex MA

Pittsfield RRF Berkshire MA 

Wheelabrator Millbury Worcester MA

Harford County WTE Fac. Harford MD

Southwest RRF (RESCO) Independent City MD

Pulaski Independent City MD

Biddeford Biddeford ME

Mid Maine Waste Action Corp. Androscoggin ME 

Penobscot Energy Recovery Co. Penobscot ME

Maine Energy Recovery York ME

Jackson Co. RRF Jackson MI

Kent Co. WTE Fac. Kent MI

Clinton Township Macomb MI

Greater Detroit RRF Wayne MI

Central Wayne Co. Sanitation Wayne MI

Auth

Elk river FFR Anoka MN

Wilmarth Plant Blue Earth & Nicollet MN

Pope-Douglas Solid Waste Douglas MN

Ramsey-Washington Goodhue MN

Red Wing Solid Waste Boiler Goodhue MN

Facility
Table A-8 (continued)

Existing MWC Facilities

Facility County State

A-12

Hennepin Energy Recovery Hennepin MN

Facility

Olmstead WTE Facility Olmstead MN

Perham Renewable RF Otter Tail MN

Fergus Falls Otter Tail MN

Polk Co. Solid Waste Resource Polk MN

Recovery

Richards Asphalt Co. Facility Savage MN

Western Lake Superior Sanitary St. Louis MN

District

Pascagoula Energy Recovery Jackson MS

Facility

Livingston (Park County) Livingston MT

NIEHS Durham NC

University City RRF Mecklenburg NC

New Hanover Co. WTE New Hanover NC

Wheelabrator Concord Merrimack NH

Lamprey Regional SW Coop. Strafford NH

SES Claremont RRF Sullivan NH 

Fort Dix RRF Burlington NJ

Camden RRF Camden NJ

Essex Co. RRF Essex NJ

Gloucester County Gloucester NJ

Union Co. RRF Union NJ

Warren Energy RF Warren NJ

Dutchess Co. RRF Dutchess NY 

Kodak RRF Monroe NY

Hempstead Nassau NY

Long Beach RRF Nassau NY 

Niagara Falls RDF WTE Niagara NY

Oneida Co. ERF Oneida NY

Onondada Co. RRF Onondaga NY

Oswgo Co. WTE Oswego NY
Table A-8 (continued)

Existing MWC Facilities

Facility County State

A-13

Babylon RRF Suffolk NY

Huntington RRF Suffolk NY

MacArthur WTE Suffolk NY

Adrirondack RRF Washington NY

Westchester RESCO Westchester NY

Montgomery Co. North RRF Montgomery OH

Montgomery Co. South RRF Montgomery OH

Miami RRF Ottawa OK

Walter B. Hall RRF Tulsa & Osage OK

Coos Bay Incinerator Coquille OR

Marion Co. WTE Marion OR

Wheelabrator Falls RRF Bucks PA

Harrisbury WTE Dauphin PA

Delaware Co. RRF Delaware PA

Lancaster Co. RRF Lancaster PA

Montgomery Co. RRF Montgomery PA

Westmoreland WTE Fac. Westmoreland PA

York Co. RR Center York PA

Foster Wheeler Charleston RR Charleston SC

Chamber Medical Tech. Of SC Hampton SC

Nashville Thermal Transfer Corp Davidson TN

Resource Authority in Sumner Co. Sumner TN

City of Cleburne Johnson TX

Panola Co. WTE Panola TX

Center RRF Shelby TX

Davis Co. WTE Davis UT

Alexandria/Arlington RRF Alexandria VA

Arlington-Pentagon Arlington VA

I-95 Energy RRF (Fairfax) Fairfax VA 

Norfolk Navy Yard Independent City VA
Table A-8 (continued)

Existing MWC Facilities

Facility County State

A-14

NASA Refuse-fired Steam Independent City VA

Generator

Harrisonburg RRF Rockingham VA

Tacoma Pierce WA

Skagit Co. RRF Skagit WA

Spokane Regional Disposal Fac. Spokane WA

Recomp Bellingham RRF Whatcom WA

Barron Co. WTE Fac. Barron WI

La Crosse Co. La Crosse WI

Sheboygan Sheboygan WI

St. Croix Co. WTE Fac. St. Croix WI

Source: U.S. Environmental Protection Agency, 1995. Municipal Waste Combustors: Background Information

Document for Promulgated Standards and Guidelines -- Public Comments and Responses.EPA-453/R95-0136. Research Triangle Park, NC. October 1995.
A-15

Table A-9

Mercury Emissions From MWCs by Combustor Type For 1995

Combustor Gas concentration, Value, capacity

type Control? ug/dscm @ 7% O Btu/lb factor

Acid Average mercury Heating Average

2

Annual Emissions

Mg/yr Tons/yr

Mass Burn Y 205 4500 0.91 12.1 13.3

N 340 4500 0.91 9.8 10.8

Refused

Derived

Fuel

Y 35 5500 0.91 0.9 1.0

N 260 5500 0.91 2.6 2.9

Modular /

Starved Air

Y 205 4500 0.74 0.0 0.0

N 340 4500 0.74 1.1 1.2

Modular /

Excess Air

Y 205 5500 0.74 0.1 0.1

N 340 5500 0.74 0.3 0.3

Total 26.9 29.6

Basis of Input Data for EPA's Emissions Calculations

1. The following criteria were used for assigning the average mercury concentrations associated with different combustor

types: 

 any non-RDF (refused derived fuel) combustor without acid gas control was assigned 340 μg/dscm @ 7

% O .2

 any non-RDF combustor with acid gas control was assigned 205 μg/dscm @ 7 % O .2

 any RDF combustor without acid gas control was assigned 260 μg/dscm @ 7 % O .2

 any RDF combustor with acid gas control was assigned 35 μg/dscm @ 7 % O .2

2. The F-factor for municipal waste combustors was assumed to be 9,570 dscf/MMBtu at 0 percent oxygen and the

heating values were assumed to be 4,500 Btu/lb for unprocessed MSW and 5,500 Btu/lb for RDF.

3. The average capacity factor, which represents the percentage of operational time a plant would operate during a year at

100 percent capacity, for modular / starved air combustors was assumed to be 74 percent. The value for all other types

of combustors was assumed to be 91 percent.

Calculations

Volumetric Flow Factor (V)

V = F-factor * heating value * (2000 lb/ton) * (20.9 / (20.9 - 7)) / (35.31 dscf/dscm) / (10 Btu/ MMBtu) 6

 For non-RDF combustors: V = 3,670 dscm @ 7% O / ton MSW 2

 For RDF combustors: V = 4,457 dscm @ 7% O / ton MSW 2

Annual Mercury Emissions (E)

E = C * V * T * CF / 1012

where:

 C = flue gas mercury concentration (μg/dscm @ 7% O )2

 V = volumetric flow factor (dscm @ 7% O / ton waste) 2

 T = MWC unit capacity (tons/year), and

CF = capacity factor (unitless)

Source: Locating and Estimating Air Emissions from Sources of Mercury and Mercury Compounds. U.S. Environmental

Protection Agency. May 1997
A-16

Table A-10

MWI Population By State

State No.

Alabama 54

Alaska 0

Arizona 14

Arkansas 39

California 23

Colorado 39

Connecticut 25

Delaware 8

District of Columbia 3

Florida 44

Georgia 103

Hawaii 0

Idaho 12

Illinois 108

Indiana 92

Iowa 34

Kansas 114

Kentucky 37

Louisiana 92

Maine 36

Maryland 82

Massachusetts 109

Michigan 287

Minnesota 119

Mississippi 21

Missouri 59

Montana 5

Nebraska 33

Nevada 0

New Hampshire 17

New Jersey 61

New Mexico 0

New York 18

North Carolina 90

North Dakota 76

Ohio 126

Oklahoma 32

Oregon 0

Pennsylvania 72

Rhode Island 11

South Carolina 26
Table A-10 (continued)

MWI Population By State

State No.

A-17

South Dakota 0

Tennessee 57

Texas 63

Utah 2

Vermont 3

Virginia 65

Washington 31

West Virginia 14

Wisconsin 10

Wyoming 7

Grand Total 2373

Source: Docket #A-91-61, Item IV-A-007.
APPENDIX B

MERCURY REMOVAL CAPABILITIES OF

PARTICULATE MATTER AND ACID GAS CONTROLS FOR UTILITIES
B-1

APPENDIX B

MERCURY REMOVAL CAPABILITIES OF PARTICULATE

MATTER AND ACID GAS CONTROLS FOR UTILITIES

Existing air pollution control devices (APCDs) on utilities typically control either particulate

matter (PM) or sulfur dioxide (SO ) emissions, or both. Nitrogen oxides may be controlled by an APCD, 2

but are usually controlled by combustion modification. Generally, a wet scrubber is used to control SO2

emissions only, while a dry scrubber can control SO emissions and PM because it is usually built with a 2

downstream PM collector. Devices that control PM only include fabric filters (FFs), electrostatic

precipitators (ESPs), mechanical collectors (cyclones), and venturi scrubbers.

Mercury, however, is not well controlled by particulate matter APCDs because mercury is

emitted as a mixture of solid and gaseous forms.

Mercury removal effectiveness is shown in this appendix as percent removal. Percent removal is

equivalent to one minus the emission modification factor (EMF). For example, a 17.3 percent removal

indicates an EMF of 0.827 or that 17.3 percent of the total mercury has been collected by that type of

control device. Calculation of EMF's is described in Section 4.1.1.3. The EMF values are presented in

Appendix C.

B.1 Scrubbers

Wet scrubbers or flue gas desulfurization (FGD) units for coal-fired plants are typically used to

remove acid gases (mainly SO emissions). Most utility boilers are equipped with an ESP or FF before 2

the wet FGD units to collect PM.

Figure B-1 shows the relationship between mercury removal and the inlet temperature for wet

FGD devices. Table B-1 summarizes available test data for FGD units. FGDs have a median mercury

removal efficiency of about 22.6 percent, with a range from 0 percent to 61.7 percent removal. The

correlation between FGD inlet temperature and mercury removal is difficult to determine. This difficulty

is compounded by having only five data sites and two of the five test sites employ flue gas bypasses in

their design. A bypass means that part of the flue gas is diverted around the FGD while the majority of

the flue gas is treated.

B-3

Table B-1

Test Data for FGD Units

Unit Control Device Hg Removal Reference

%

EPRI Site 11 Wet limestone FGD (inlet Hg

concentration of 9.9 g/dscm) 10.87 Radian, 1993a

EPRI Site 12 Wet limestone FGD 0.00 Radian, 1993b

NSP Sherburne 1 & 2 Test A Wet limestone FGD (inlet Hg

concentration of 8.1 g/dscm) 22.63 Interpoll, 1990a

NSP Sherburne 1 & 2 Test B Wet limestone FGD (inlet Hg

concentration of 11.6 g/dscm) 59.3 Interpoll, 1991

DOE Yates Wet limestone and jet bubbling

reactor FGD (inlet Hg concentration 45.91 EPRI, 1993a

of 6.0 g/dscm)

DOE Coal Creek Wet lime FGD (inlet Hg

concentration of 10.0 g/dscm) 12.05 Battelle, 1993a

EPRI Site 20 Wet limestone FGD (inlet Hg

concentration of 12.5 g/dscm) 20.15 Radian, 1994b

EPRI Site 101 Wet lime FGD (inlet Hg

concentration of 5.6 g/dscm) 61.67 Radian, 1994c

DOE Paradise Wet limestone FGD (inlet Hg Southern

concentration of 9.5 g/dscm) 45.10 Research

Institute, 1995a

Median 22.63

Mean 30.85

Standard deviation 22.57

 This unit was re-tested for mercury as part of a ESP/FGD system. Since there was no way of determining which a

component (the ESP or the FGD) was responsible for any mercury removal, the ESP was given the full credit for

removal, as shown in the site 12 ESP data in Table B-4.

B.2 SDA or Dry Scrubbing

A spray dryer adsorber (SDA) process is a dry scrubbing system followed by a particulate

control device. A lime/water slurry is sprayed into the flue gas stream and the resulting dry solids are

collected by an ESP or an FF. 

Figure B-2 shows the relationship between mercury removal and the inlet temperature for the

SDA/FF systems. Available SDA data are presented in Table B-2. SDA/FF systems have a median

mercury removal efficiency of about 23.9 percent, with a range from 0 percent to 54.5 percent removal.

B-5

Table B-2

Spray Dryer Adsorption Data

Unit Control Device Hg Removal % Reference

EPRI Site 14 SDA/FF (inlet Hg concentration of 1.0 0 Radian, 1993c

g/dscm)

DOE Springerville SDA/FF (inlet Hg concentration of 8.3 2.16 Southern Research

g/dscm) Institute, 1993a

Sherburne 3 Test A SDA/FF (inlet Hg concentration of 6.8 45.71 Interpoll, 1990b

g/dscm)

Sherburne 3 Test B SDA/FF (inlet Hg concentration of 13.4 54.5 Interpoll, 1991

g/dscm)

Median 23.94

Mean 25.59

Standard deviation 28.54

B.3 Fabric Filters

Figure B-3 shows the relationship between mercury removal and the PM collection efficiency

(percent) for FFs (controlling coal-fired units). Available FF data are presented in Table B-3. Fabric

filters have a median mercury removal efficiency of about 8.39 percent, with a range from 0 percent to

73.36 percent removal.

B-7

Table B-3

Fabric Filter Data

Unit Control Device Hg Removal % Reference

EPRI Site 13 FF (inlet Hg concentration of 0 Radian, 1993d

0.3 g/dscm)

EPRI Site 115 FF (inlet Hg concentration of 73.36 Carnot, 1994a

1.8 g/dscm)

NSP Riverside 6 & 7 FF (inlet Hg concentration of 0 Interpoll, 1992a

4.8 g/dscm)

DOE Niles #2 w/NO FF (inlet Hg concentration of 8.39 Battelle, 1993b X 

25.8 g/dscm)

DOE Boswell FF (inlet Hg concentration of 60.59 Weston, 1993a

6.4 g/dscm)

Median 8.39

Mean 28.47

Standard deviation 35.61

B.4 Electrostatic Precipitators

Electrostatic precipitators are the most widely used control device by the fossil fuel-fired electric

utility industry. There are two design locations for ESPs, cold-side (CS) and hot-side (HS). Cold-side

ESPs are located after the air preheater, thus it is subjected to a lower flue gas temperature than a hotside ESP which is located before the air preheater.

Figure B-4 shows the relationship between mercury removal and the PM collection efficiency

(percent) for cold-side ESPs (controlling coal-fired units). Table B-4 presents available test data for such

EPSs. Cold-side ESPs have a median mercury removal efficiency of about 16.2 percent, with a range

from 0 percent to 82.4 percent removal.

Figure B-5 shows the relationship between mercury removal and the PM collection efficiency

(percent) for hot-side ESPs (controlling coal-fired units). Available test data for hot-side ESPs

(controlling coal-fired units) are shown in Table B-5. There was no apparent control of mercury by a

hot-side ESP. However, the data were collected from only one emission test where two separate sample

runs were analyzed.

Figure B-6 shows the relationship between mercury removal and the PM collection efficiency

(percent) for cold-side ESPs (controlling oil-fired units). Table B-6 presents available test data for such

configurations. In these emission tests cold-side ESPs (controlling oil-fired units) had a median mercury

removal efficiency of about 62.4 percent, with a range from 41.7 percent to 83 percent removal. It

should be noted that data for mercury control by cold-side ESPs (controlling oil-fired units) were

available from only two test sites.

B-9

Table B-4

Test Data for Cold-Side Electrostatic Precipitators (Controlling Coal-Fired Units)

Unit Control Device Hg Reference

Removal

%

EPRI Site 11 ESP, CS (inlet Hg concentration of 3.4 g/dscm) 0.00 Radian, 1993a

EPRI Site 12 ESP, CS (inlet Hg concentration of 9.1 g/dscm) 82.35 Radian, 1993b

EPRI Site 15 ESP, CS (inlet Hg concentration of 4.9 g/dscm) 0.00 Radian, 1992a

EPRI Site 102 ESP, CS (inlet Hg concentration of 9.0 g/dscm) 0.00 Radian, 1993e

NSP High Bridge 3,4,5,6 ESP, CS (inlet Hg concentration of 4.4 g/dscm) 6.87 Interpoll,

1992b

NSP High Bridge 1,3,4 ESP, CS (inlet Hg concentration of 5.1 g/dscm) 8.21 Interpoll,

1992c

NSP Black Dog #2 ESP, CS (inlet Hg concentration of 2.8 g/dscm) 21.56 Interpoll,

1992d

NSP Riverside #8 ESP, CS (inlet Hg concentration of 2.9 g/dscm) 0.00 Interpoll,

1992e

EPRI Site 114 / Test A ESP, CS (inlet Hg concentration of 10.6 g/dscm) 29.8 Radian, 1994a

EPRI Site 114 / Test B ESP, CS (inlet Hg concentration of 10.6 g/dscm) 16.16 Radian, 1994a

DOE Niles #2 ESP, CS (inlet Hg concentration of 24.7 g/dscm) 26.55 Battelle, 1993c

DOE Yates ESP, CS (inlet Hg concentration of 5.9 g/dscm) 55.23 EPRI, 1993a

DOE Coal Creek ESP, CS (inlet Hg concentration of 11.0 g/dscm) 13.15 Battelle, 1993a

EPRI Site 16/OFA/LNO Burners ESP, CS (inlet Hg concentration of 11.5 X g/dscm) 54.8 EPRI, 1993b

EPRI Site 16/OFA ESP, CS (inlet Hg concentration of 7.6 g/dscm) 9.38 EPRI, 1993b

DOE Cardinal ESP, CS (inlet Hg concentration of 2.3 g/dscm) 73.9 EERC, 1993

DOE Baldwin ESP, CS (inlet Hg concentration of 7.0 g/dscm) 26.13 Weston, 1993b

EPRI Site 116 ESP, CS (inlet Hg concentration of 12.1 g/dscm) 7.59 Radian, 1994d

Median 14.66

Mean 23.98

Standard deviation 25.87


B-12

Table B-6

Test Data for Cold-Side Electrostatic Precipitators (Controlling Oil-Fired Units)

Unit Control Device Hg Removal % Reference

EPRI Site 112 (oil-fired) ESP, CS (inlet Hg concentration of 1.8 83 Carnot, 1994b

g/dscm)

EPRI Site 118 (oil-fired) ESP, CS (inlet Hg concentration of 1.4 41.7 Carnot, 1994c

g/dscm)

Median 62.35

Mean 62.35

Standard deviation 29.2

B.5 Mechanical Collectors and Venturi Scrubbers

Mechanical collectors typically have very low collection efficiencies, often lower than 30

percent for particles in the 0 to 0.3 μm size range. These devices are used as gross particulate removal

devices before ESPs or as APCDs on oil-fired units. Venturi scrubbers can be effective for particulate

control but require high pressure drops (more than 50 or 60 in. of water) for small particles. Even with

high pressure drops, ESPs and FFs are normally more effective for submicron particles. Mechanical

collectors and venturi scrubbers are not expected to provide effective mercury removal, especially for

those mercury compounds concentrated in the submicron PM fractions and in the vapor phase and, thus,

are not discussed in this study.

B.6 References for Appendix B

Battelle, 1993a. Preliminary draft emissions report for Coal Creek Station - Unit 2 (Cooperative Power

Association) for the Comprehensive Assessment of Toxic Emissions from Coal-Fired Power Plants. 

Prepared for the Department of Energy/Pittsburgh Energy Technology Center (DOE/PETC). DOE

contract # DE-AC22-93PC93251. December 1993.

Battelle, 1993b. Preliminary draft emissions report for Niles Station Boiler No. 2 with SNOX (Ohio

Edison) for the Comprehensive Assessment of Toxic Emissions from Coal-Fired Power Plants. Prepared

for the Department of Energy/Pittsburgh Energy Technology Center (DOE/PETC). DOE contract # DEAC22-93PC93251. December 1993.

Battelle, 1993c. Preliminary draft emissions report for Niles Station Boiler No. 2 (Ohio Edison) for the

Comprehensive Assessment of Toxic Emissions from Coal-Fired Power Plants. Prepared for the

Department of Energy/Pittsburgh Energy Technology Center (DOE/PETC). DOE contract # DE-AC22-

93PC93251. December 1993.

Carnot, 1994a. Preliminary draft emissions report for EPRI Site 115, Field Chemical Emissions

Monitoring Project. Prepared for Electric Power Research Institute. Carnot report No. EPRIE10106/R022C855.T. November 1994.
B-13

Carnot, 1994b. Preliminary draft emissions report for EPRI Site 112, Field Chemical Emissions

Monitoring Project. Prepared for Electric Power Research Institute. Carnot report No. EPRIE10106/R016C374.T. March 1994.

Carnot, 1994c. Preliminary draft emissions report for EPRI Site 118, Field Chemical Emissions

Monitoring Project. Prepared for Electric Power Research Institute. Carnot report No. EPRIE10106/R140C928.T. January 1994.

EERC, Inc., 1993. Preliminary draft emissions report for Cardinal Station - Unit 1 (American Electric

Power) for the Comprehensive Assessment of Toxic Emissions from Coal-Fired Power Plants. Prepared

for the Department of Energy/Pittsburgh Energy Technology Center (DOE/PETC). DOE contract # DEAC22-93PC93252. December 1993.

EPRI, 1993a. Preliminary draft emissions report for Plant Yates Unit No. 1 (Georgia Power Company)

for the Comprehensive Assessment of Toxic Emissions from Coal-Fired Power Plants. Prepared for the

Department of Energy/Pittsburgh Energy Technology Center (DOE/PETC). EPRI Report No. DCN 93-

643-004-03. December 1993.

EPRI, 1993b. Preliminary draft emissions report for EPRI Site 16 (OFA and OFA/Low NOx) for the

Clean Coal Technology Project (CCT). Prepared for the Department of Energy/Pittsburgh Energy

Technology Center (DOE/PETC). EPRI report No. DCN 93-209-061-01. November 1993.

Interpoll Laboratories, Inc., 1990a. Results of the May 1, 1990 Trace Metal Characterization Study on

Units 1 & 2 at the Northern States Power Company Sherburne Plant. Prepared for Northern States Power

Company. Report No. 0-3033E. July 1990.

Interpoll Laboratories, Inc., 1990b. Results of the March 27, 1990 Trace Metal Characterization Study

on Unit 3 at the Northern States Power Company Sherburne Plant. Prepared for Northern States Power

Company. Report No. 0-3005. June 1990.

Interpoll Laboratories, Inc., 1991. Results of the September 10 and 11, 1991 Mercury Removal Tests on

Units 1 & 2, and Unit 3 Scrubber Systems at the Northern States Power Company Sherburne Plant. 

Prepared for Northern States Power Company. Report No. 1-3409. October 1991.

Interpoll Laboratories, 1992a. Results of the Air Toxic Emission Study on the No. 6 & 7 Boilers at the

Northern States Power Company Riverside Plant. Prepared for Northern States Power Company. Report

No. 1-3468A. February 1992.

Interpoll Laboratories, Inc., 1992b. Results of the November 7, 1991 Air Toxic Emission Study on the

No. 3, 4, 5 & 6 Boilers at the Northern States Power Company High Bridge Plant. Prepared for Northern

States Power Company. Report No. 

1-3453. January 1992.

Interpoll Laboratories, Inc., 1992c. Results of the November 5, 1991 Air Toxic Emission Study on the

No. 1, 3, & 4 Boilers at the Northern States Power Company Black Dog Plant. Prepared for Northern

States Power Company. Report No. 1-3451. January 1992.
B-14

Interpoll Laboratories, Inc., 1992d. Results of the January 1992 Air Toxic Emission Study on the No. 2

Boiler at the Northern States Power Company Black Dog Plant. Prepared for Northern States Power

Company. Report No. 2-3496. May 1992.

Interpoll Laboratories, Inc., 1992e. Results of the July 1992 Air Toxic Emission Study on the No. 8

Boiler at the Northern States Power Company Riverside Plant. Prepared for Northern States Power

Company. Report No. 2-3590. September 1992.

Radian Corp., 1992a. Preliminary draft emissions report for EPRI Site 15, Field Chemical Emissions

Monitoring Project. Prepared for Electric Power Research Institute. EPRI report No. DCN 93-213-152-

26. October 1992.

Radian Corp., 1993a. Preliminary draft emissions report (and mercury retest) for EPRI Site 11, Field

Chemical Emissions Monitoring Project. Prepared for Electric Power Research Institute. EPRI report

Nos. DCN 92-213-152-24 and DCN 92-213-152-48. November 1992/October 1993.

Radian Corp., 1993b. Preliminary draft emissions report (and mercury retest) for EPRI Site 12, Field

Chemical Emissions Monitoring Project. Prepared for Electric Power Research Institute. EPRI report

Nos. DCN 92-213-152-27 and DCN 93-213-152-49. November 1992/October 1993.

Radian Corp., 1993c. Preliminary draft emissions report for EPRI Site 14, Field Chemical Emissions

Monitoring Project. Prepared for Electric Power Research Institute. EPRI report No. DCN 93-213-152-

28. November 1992.

Radian Corp., 1993d. Preliminary draft emissions report for EPRI Site 13, Field Chemical Emissions

Monitoring Project. Prepared for Electric Power Research Institute. EPRI report No. DCN 93-213-152-

36. February 1993.

Radian Corp., 1993e. Preliminary draft emissions report for EPRI Site 102, Field Chemical Emissions

Monitoring Project. Prepared for Electric Power Research Institute. EPRI report No. DCN 92-213-152-

35. February 1993.

Radian Corp., 1994a. Preliminary draft emissions report for EPRI Site 114, Field Chemical Emissions

Monitoring Project. Prepared for Electric Power Research Institute. EPRI report No. DCN 92-213-152-

51. May 1994.

Radian Corp., 1994b. Preliminary draft emissions report for EPRI Site 20, Field Chemical Emissions

Monitoring Project. Prepared for Electric Power Research Institute. EPRI report No. DCN 92-213-152-

51. March 1994.

Radian Corp., 1994c. Preliminary draft emissions report for EPRI Site 101, Field Chemical Emissions

Monitoring Project. Prepared for Electric Power Research Institute. EPRI report No. DCN 94-643-015-

02. October 1994.

Radian Corp., 1994d. Preliminary draft emissions report for EPRI Site 116, Field Chemical Emissions

Monitoring Project. Prepared for Electric Power Research Institute. EPRI report No. DCN 94-213-152-

55. October 1994.
B-15

Southern Research Institute, 1993a. Preliminary draft emissions report for Springerville Generating

Station Unit No. 2 (Tucson Electric Power Company) for the Comprehensive Assessment of Toxic

Emissions from Coal-Fired Power Plants. Prepared for the Department of Energy/Pittsburgh Energy

Technology Center (DOE/PETC). DOE contract # DE-AC22-93PC93254. SRI Report No. SRI-ENV93-1049-7960. December 1993.

Southern Research Institute, 1993b. Preliminary draft emissions report for EPRI Site 110 (baseline and

with NOx control), Field Chemical Emissions Monitoring Project. Prepared for Southern Company

Services. Report No. SRI-ENV-92-796-7496. October 1993.

Southern Research Institute, 1995a. Preliminary draft emissions report for Paradise Generating Station

Unit No. 1 (Tennessee Valley Authority) for the Comprehensive Assessment of Toxic Emissions from

Coal-Fired Power Plants. Prepared for the Department of Energy/Pittsburgh Energy Technology Center

(DOE/PETC). DOE contract # DE-AC22-93PC93255. SRI Report No. SRI-ENV-95-338-7960. May

1995.

Roy F. Weston, Inc., 1993a. Preliminary draft emissions report for Boswell Energy Center - Unit 2

(Minnesota Power Company) for the Comprehensive Assessment of Toxic Emissions from Coal-Fired

Power Plants. Prepared for the Department of Energy/Pittsburgh Energy Technology Center

(DOE/PETC). DOE contract # DE-AC22-93PC93255. Weston project # 10016-011, Weston report #

DOE017G.RP1. December 1993.

Roy F. Weston, Inc., 1993b. Preliminary draft emissions report for Baldwin Power Station - Unit 2

(Illinois Power Company) for the Comprehensive Assessment of Toxic Emissions from Coal-Fired

Power Plants. Prepared for the Department of Energy/Pittsburgh Energy Technology Center

(DOE/PETC). DOE contract # DE-AC22-93PC93255. Weston project # 10016-011, Weston report #

DOE018G.RP1. December 1993.
APPENDIX C

EMISSION MODIFICATION FACTORS FOR

UTILITY BOILER EMISSION ESTIMATES
C-1

Table C-1

Emission Modification Factors for Utility Boiler Emission Estimates a

Type of APCD or Boiler EMF Factor

Fabric Filter 0.626

Spray Dryer Adsorber (includes a fabric filter) 0.701

Electrostatic precipitator (cold-side) 0.696

Electrostatic precipitator (hot-side) 1.000

Electrostatic precipitator (oil-fired unit) 0.315

Particulate matter scrubber 0.957

Fluidized gas desulfurization scrubber 0.656

Circulating fluidized bed combustor 1.000

Cyclone-fired boiler without NOx control (wet bottom, coal-fired) 0.856

Front-fired boiler without NOx control (dry bottom, coal-fired) 0.764

Front-fired boiler without NOx control (dry bottom, gas-fired) 1.000

Tangential-fired boiler without NOx control (before a hot-side ESP, 1.000

coal-fired)

Tangential-fired boiler with NOx control (before a hot-side ESP, coal- 0.748

fired)

Front-fired boiler without NOx control (dry bottom, oil-fired) 1.000

Front-fired boiler with NOx control (dry bottom, oil-fired) 1.000

Opposed-fired boiler without NOx control (dry bottom oil-fired) 0.040

Tangentially-fired boiler without NOx control (dry bottom, oil-fired) 1.000

Tangentially-fired boiler with NOx control (dry bottom, oil-fired) 1.000

Opposed-fired boiler with NOx control (dry bottom, coal-fired) 0.812

Opposed-fired boiler without NOx control (wet bottom, coal-fired) 0.918

Tangentially-fired boiler without NOx control (dry bottom, coal-fired) 1.000

Tangentially-fired boiler with NOx control (dry bottom, coal-fired) 0.625

Vertically-fired boiler with NOx control (dry bottom, coal-fired) 0.785

 To calculate mercury control efficiency for a specific boiler/control device configuration, the EMF is subtracted from 1. a