United States EPA-452/R-96-001b (/
Environmental Protection June 1996
Agency
Air
Mercury Study
Report to Congress
Volume II:
An Inventory of Anthropogenic
Mercury Emissions in the
United States
SAB REVIEW DRAFT
U S Environmental Protection Agency
Region 5, Library (PL-12J)
77 West Jackson Boulevard, 12th Hoof
Chicago, It 60604-3590
Office of Air Quality Planning & Standards
and
Office of Research and Development
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MERCURY STUDY REPORT TO CONGRESS
VOLUME H:
AN INVENTORY OF ANTHROPOGENIC MERCURY
EMISSIONS IN THE UNITED STATES
SAB REVIEW DRAFT
June 1996
U.S. Environmental Protection Agency
Region 5.Library (PL- 12J)
77 West Jackson Boulevard, 12th Floor
Chicago, IL 60604-3590
Office of Air Quality Planning and Standards
and
Office of Research and Development
U.S. Environmental Protection Agency
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TABLE OF CONTENTS
Page
U.S. EPA AUTHORS iii
SCIENTIFIC PEER REVIEWERS iv
WORK GROUP AND U.S. EPA/ORD REVIEWERS v
xi LIST OF TABLES vi
K LIST OF FIGURES ix
>r LIST OF SYMBOLS, UNITS AND ACRONYMS x
Vr,
^t EXECUTIVE SUMMARY I
=&
1. INTRODUCTION 1-1
^ 1.1 Overview of Sources '. 1-1
V 1.2 Study Approach and Uncertainties 1-2
1.3 Organization of the Rest of the Document 1-4
2. MERCURY TRENDS 2-1
2.1 Trends in the Atmospheric Mercury Burden 2-1
2.2 Trends in Mercury Consumption 2-4
3. ANTHROPOGENIC AREA SOURCES OF MERCURY EMISSIONS 3-1
3.1 Electric Lamp Breakage 3-1
3.2 General Laboratory Use 3-6
3.3 Dental Preparation and Use 3-7
3.4 Mobile Sources 3-7
3.5 Paint Use 3-8
3.6 Agricultural Burning 3-8
3.7 Other Area Sources 3-9
4. ANTHROPOGENIC POINT SOURCES OF MERCURY EMISSIONS 4-1
4.1 Combustion Sources 4-1
4.1.1 Medical Waste Incinerators 4-2
4.1.2 Municipal Waste Combustors 4-5
4.1.3 Utility Boilers 4-17
4.1.4 Commercial/Industrial Boilers 4-23
4.1.5 Residential Boilers 4-28
4.1.6 Sewage Sludge Incinerators 4-29
4.1.7 Crematories 4-30
4.1.8 Wood Combustion 4-30
4.1.9 Hazardous Waste Combustors 4-31
4.2 Manufacturing Sources 4-33
4.2.1 Primary Lead Smelting 4-33
4.2.2 Secondary Mercury Production 4-37
4.2.3 Chlor-alkali Production Using the Mercury Cell Process 4-39
4.2.4 Cement Manufacturing 4-43
4.2.5 Primary Copper Smelting 4-45
4.2.6 Lime Manufacturing 4-48
4.2.7 Electrical Apparatus Manufacturing 4-51
June 1996 i SAB REVIEW DRAFT
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TABLE OF CONTENTS (continued)
4.2.8 Instrument (Thermometers) Manufacturing . 4-55
4.2.9 Carbon Black Production 4-56
4.2.10 Battery Production 4-57
4.2.11 Primary Mercury Production 4-62
4.2.12 Mercury Compounds Production 4-64
4.2.13 Byproduct Coke Production 4-64
4.2.14 Petroleum Refining 4-68
4.3 Miscellaneous Sources 4-69
4.3.1 Geothermal Power Plants 4-69
4.3.2 Pigments, Oil Shale Retorting, Mercury Catalysts, Turf Products and
Explosives 4-69
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-l
APPENDIX B MERCURY REMOVAL CAPABILITIES OF PARTICULATE
MATTER AND ACID GAS CONTROLS FOR UTILITIES B-l
APPENDIX C EMISSION MODIFICATION FACTORS FOR UTILITY BOILER
EMISSION ESTIMATES . C-l
June 1996 . ii SAB REVIEW DRAFT
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U.S. EPA AUTHORS
€
Principal Author:
Martha H. Keating
Office of Air Quality Planning and Standards
Research Triangle Park, NC
Contributing Authors:
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
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
June 1996' iii SAB REVIEW DRAFT
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SCIENTIFIC PEER REVIEWERS
Brian J. Allee, Ph.D.
Harza Northwest, Incorporated
Thomas D. Atkeson, Ph.D.
Florida Department of Environmental
Protection
Steven M. Bartell, Ph.D.
SENES Oak Ridge, Inc.
Mike Bolger, Ph.D.
U.S. Food and Drug Administration
James P. Butler, Ph.D.
University of Chicago
Argonne National Laboratory
Rick Canady, Ph.D.
Agency for Toxic Substances and Disease
Registry
Rufus Chaney, Ph.D.
U.S. Department of Agriculture
Tim Eder
Great Lakes Natural Resource Center
National Wildlife Federation for the
States of Michigan and Ohio
William F. Fitzgerald, Ph.D.
University of Connecticut
Avery Point
Robert Goyer, Ph.D.
National Institute of Environmental Health
Sciences
George Gray, Ph.D.
Harvard School of Public Health
Terry Haines, Ph.D.
National Biological Service
Joann L. Held
New Jersey Department of Environmental
Protection & Energy
Gerald J. Keeler, Ph.D.
University of Michigan
Ann Arbor
Leonard Levin, Ph.D.
Electric Power Research Institute
Malcom Meaburn, Ph.D.
National Oceanic and Atmospheric
Administration
U.S. Department of Commerce
Paul Mushak, Ph.D.
PB Associates
Jozef M. Pacyna, Ph.D.
Norwegian Institute for Air Research
Ruth Patterson, Ph.D.
Cancer Prevention Research Program
Fred Gutchinson Cancer Research Center
Donald Porcella, Ph.D.
Electric Power Research Institute
Charles Schmidt
U.S. Department of Energy
Pamela Shubat, Ph.D.
Minnesota Department of Health
Alan H. Stern, Dr.P.H.
New Jersey Department of Environmental
Protection & Energy
Edward B. Swain, Ph.D.
Minnesota Pollution Control Agency
M. Anthony Verity, M.D.
University of California
Los Angeles
June 1996
IV
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WORK GROUP AND U.S. EPA/ORD REVIEWERS
Core Work Group Reviewers:
Dan Axelrad, U.S. EPA
Office of Policy, Planning and Evaluation
Angela Bandemehr, U.S. EPA
Region 5
Jim Darr, U.S. EPA
Office of Pollution Prevention and Toxic
Substances
Thomas Gentile, State of New York
Department of Environmental Conservation
Arnie Kuzmack, U.S. EPA
Office of Water
David Layland, U.S. EPA
Office of Solid Waste and Emergency
Response
Karen Levy, U.S. EPA
Office of Policy Analysis and Review
Steve Levy, U.S. EPA
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
U.S. EPA/ORD Reviewers:
Robert Beliles, Ph.D., D.A.B.T.
National Center for Environmental Assessment
Washington, DC
Eletha Brady-Roberts
National Center for Environmental Assessment
Cincinnati, OH
Annie M. Jarabek
National Center for Environmental Assessment
Research Triangle Park, NC
Matthew Lorber
National Center for Environmental Assessment
Washington, DC
Susan Braen Norton
National Center for Environmental Assessment
Washington, DC
Terry Harvey, D.V.M.
National Center for Environmental Assessment
Cincinnati, OH
June 1996
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LIST OF TABLES
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 National Mercury Emission Rates by Category ES-6
ES-4 Best Point Estimates of Mercury Emissions from Anthropogenic
Sources: 1990-1993 ES-7
ES-5 Best Point Estimates of Mercury Emissions from Anthropogenic Area
Sources: 1990-1993 ES-8
ES-6 Best Point Estimates of Mercury Emissions from Anthropogenic
Combustion Point Sources: 1990-1993 ES-9
ES-7 Best Point Estimates of Mercury Emissions from Anthropogenic
Manufacturing Sources: 1990-1993 ES-10
ES-8 Best Point Estimates of Mercury Emissions from Miscellaneous
Anthropogenic Emission Point Sources: 1990-1993 ES-11
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 2-5
3-1 Best Point Estimates of Mercury Emissions from Anthropogenic Area
Sources: 1990-1993 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: 1990-1993 4-1
4-2 Best Point Estimates of Mercury Emissions from Anthropogenic Combustion
Point Sources: 1990-1993 4-3
4-3 Estimated Discards of Mercury in Products in Municipal Solid Waste
(in tons) 4-10
4-4 Estimated Discards of Mercury in Products in Municipal Solid Waste
(in percent) 4-11
4-5 Estimated .Discards of Mercury in Batteries 4-13
4-6 Estimated Discards of Mercury in Paint Residues 4-15
4-7 Estimated Discards of Mercury in Thermostats 4-16
4-8 Comparison of Mercury Concentrations in Raw and Cleaned Coal 4-26
4-9 Best Point Estimate of Mercury Emissions from Utility Boilers: 1990 4-27
4-10 Best Point Estimate of Mercury Emissions from Anthropogenic Manufacturing
Sources: 1990-1993 4-34
4-11 1990 U.S. Primary Lead Smelters and Refineries 4-35
4-12 1989 U.S. Mercury Recyclers 4-37
4-13 1991 U.S. Mercury-Cell Chlor-Alkali Production Facilities 4-40
4-14 1992 U.S. Primary Copper Smelters and Refineries 4-46
4-15 Mercury Ore Concentrate and Emissions from Primary Copper
Smelters in the U.S 4-48
June 1996 vi SAB REVIEW DRAFT
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LIST OF TABLES (continued)
4-16 Leading 1991 U.S. Lime Producing Plants 4-49
4-17 Discards of Mercury in Electric Switches . . 4-52
4-18 1992 U.S. Fluorescent Lamp Manufacturers' Headquarters 4-54
4-19 1992 U.S. Carbon Black Production Facilities 4-58
4-20 1992 U.S. Mercuric Oxide, Alkaline Manganese, or Zinc-
Carbon Button Cell Battery Manufacturers 4-59
4-21 Emission Source Parameters for an Integrated Mercury
Button Cell Manufacturing Facility 4-61
4-22 1992 U.S. Byproduct Mercury-Producing Gold Mines 4-62
4-23 1991 U.S. Mercury Compound Producers. 4-65
4-24 - 1991 U.S. Byproduct Coke Producers 4-66
4-25 Best Point Estimates of Mercury Emissions from Miscellaneous
Anthropogenic Emission Sources: 1990-1993 4-70
4-26 1992 U.S. Geothermal Power Plants 4-71
4-27 Mercury Emission Factors for Geothermal Power Plants 4-71
5-1 Best Point Estimates of 1990 National Mercury Emission Rates
by Category 5-2
5-2 Best Point Estimates of Mercury Emissions from Anthropogenic
Sources: 1990-1993 5-3
5-3 Mercury Area Sources Allocation Methodology 5-5
A-l Mercury Emissions From Utility Boilers, By State and Fuel Type A-l
A-2 Estimates of 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 1991 A-6
A-4 Estimates of Mercury Emissions From Oil-Fired Commercialflndustrial
Boilers On a Per-State Basis For 1991 A-7
A-5 Estimates of Coal, Natural Gas, and Oil Consumption in the Residential
Sector Per State (Trillion Btu) A-8
A-6 Estimates of Mercury Emissions From Coal-Fired Residential Boilers
on a Per-State Basis For 1991 A-9
A-7 Estimates of Mercury Emissions From Oil-Fired Residential Boilers
on a Per-State Basis For 1991 A-10
A-8 1991 U.S. Crematory Locations by State A-ll
A-9 Existing MWC Facilities (As of December, 1991) A-12
A-10 Mercury Emissions From MWCs by Combustor Type For 1994 A-20
A-ll MWI Population By State A-22
A-12 Mercury Emissions From Model Medical Waste Incinerators A-24
A-13 U.S. Sewage Sludge Incinerators A-25
June 1996 vii • SAB REVIEW DRAFT
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LIST OF TABLES (continued)
B-l Test Data for FGD Units B-3
B-2 Spray Dryer Adsorption Data B-4
B-3 Fabric Filter Data B-6
B-4 Test Data for Cold-Side Electrostatic Precipitators (Controlling
Coal-Fired Units) B-8
B-5 Test Data for Hot-Side Electrostatic Precipitators (Controlling
Coal-Fired Units) B-10
B-6 Test Data for Cold-Side Electrostatic Precipitators (Controlling
Oil-Fired Units) B-ll
C-l Emission Modification Factors for Utility Boiler Emission Estimates C-l
June 1996 viii SAB REVIEW DRAFT
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'age
LIST OF FIGURES
ES-l Annual Mercury Emissions All Sources, All Species ES-13
2-1 The Global Mercury Cycle 2-2
2-2 Comparison of Current and Pre-Industrial Mercury Budgets and Fluxes ........ 2-3
2-3 U.S. Mercury: Supply, Demand, Secondary Production . 2-6
3-1 Overall Fate of Mercury from Used Mercury-Containing
Fluorescent Lamps 3-5
4-1 Municipal Waste Combustor Facilities 4-6
4-2 Discards of Mercury in Municipal Solid Waste, 1989 4-12
4-3 Estimated Discards of Mercury in Electric Lighting in Municipal
Solid Waste 4-14
4-4 Estimated Discards of Mercury in Pigments in Municipal Solid Waste 4-17
4-5 Utility Boiler Locations 4-18
4-6 Comparison of Mercury Removal Efficiences Without Activated Carbon
Injection 4-20
4-7 Mercury Emissions from Oil- and Natural-Gas Fired Boilers 4-25
4-8 Mercury Emissions from Coal-Fired Boilers 4-25a
4-9 Primary Lead Smelters 4-35
4-10 Chlor-Alkali Production Facilities 4-41
4-11 Cement Manufacturing Plants . 4-43
4-12 Primary Copper Smelters 4-46
4-13 Coke Oven Locations 4-67
5-1 Annual Mercury Emissions — All Sources, All Species 5-6
B-l 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-5
B-4 Removal of Mercury By Electrostatic Precipitators (Cold-Side, Coal) B-7
B-5 Removal of Mercury By Electrostatic Precipitators (Hot-Side, Coal) B-9
B-6 Removal of Mercury By Electrostatic Precipitators (Oil) B-10
June 1996 ix SAB REVIEW DRAFT
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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 Circulized 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)
Ib Pound
MB/REF Mass burn/refractory wall
MB/RC Mass burn/rotary waterwall
MBAVW 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
Nm3 Normal cubic meter
NSPS New Source Performance Standard
NYCC New York City Cycle
OPP U.S. EPA Office of Pesticides Programs
OSHA Occupational Safety and Health Administration
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LIST OF SYMBOLS, UNITS AND ACRONYMS (continued)
OSW U.S. EPA Office of Solid Waste
PM Paniculate 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 intineratof
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
June 1996
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EXECUTIVE SUMMARY
Section 112(n)(l)(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. Hie 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 a seven-volume Mercury Study Report to
Congress. This volume — Volume n 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
substiration/elimination that could be used to achieve these reductions (as presented in Volume VII 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 m 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.
June 1996 ES-1 SAB REVIEW DRAFT
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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 one-third 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.1 It also requires 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. Emission factors are generated
from emission test data, from engineering analyses based on mass balance techniques, or from transfer
1 The emission factors used in developing this mercury emissions inventory are generally consistent with
those presented in the U.S. EPA document entitled Locating and Estimating Air Emissions from Sources of
Mercury and Mercury Compounds (EPA/454/R-93-023) September 1993. (Some of the nationwide emission
estimates may vary slightly between the two documents because this report uses the most recendy available data,
whereas the emission factor document mentioned above is based on a baseline year of 1990.)
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Table ES-1
Sources of Anthropogenic Mercury Emissions Examined in this Inventory
Area
Electric lamp breakage
Paints use
Laboratory use
Dental preparations
Mobile sources'
Agricultural burning2
Landfills"
Sludge application*
Point
Combustion
Utility boilers
Commercial/industrial
boilers
Residential boilers
Municipal waste
combustors
Medical waste incinerators
Sewage sludge
incinerators
Hazardous waste
combustors
Wood-fired boilers
Residential woodstoves*
Crematories
Manufacturing
Chlor-alkali production
Lime manufacturing
Primary mercury production
Mercury compounds
production3
Battery production
Electrical apparatus
manufacturing
Carbon black production
Byproduct coke production2
Primary copper smelting
Cement manufacturing
Primary lead smelting
Petroleum refining*
Instrument manufacturing
Secondary mercury
production
Zinc mining2
Fluorescent lamp recycling
Miscellaneous
Oil shale retorting
Mercury catalysts
Pigment production
Explosives
manufacturing*
Geothennal power
plants
Turf products
a Potential anthropogenic sources of mercury for which emissions were not estimated.
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of information 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. Table ES-4 shows
the distribution of estimated emissions among the four major classes of anthropogenic emission
sources (area sources, combustion point sources, manufacturing point sources;, and miscellaneous point
sources). Tables ES-S through ES-8 illustrate the distributions among individual source categories for
these four classes, along with the date of underlying data, the degree of uncertainty, and the basis for
the emission estimates. All of these emissions estimates should be regarded as best estimates given
available data.
Of the estimated 220 Megagrams (Mg) (243 tons) of mercury emitted annually into the
atmosphere by anthropogenic sources in the United States, approximately 85 percent is from
combustion point sources, 13 percent is from manufacturing point sources, 1 percent is from
miscellaneous sources and 1 percent is from area sources. Four specific source categories account for
approximately 83 percent of the total anthropogenic emissions—medical waste incineration
(27 percent), municipal waste combustion (23 percent), utility boilers (21 percent), and
commercial/industrial boilers (12 percent). It should be noted that the U.S. EiPA has finalized mercury
emission limits for municipal waste combustors, and has proposed mercury emission limits for medical
waste incinerators. These emission limits will reduce mercury emissions from these sources by 90
percent.
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 city level) of
the mercury point sources described in this document were known.
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Table ES-2
Anthropogenic Mercury Sources With Sufficient
Data to Estimate National Emissions
Area
Electric lamp breakage
Laboratory use
Dental preparation
Point
Combustion
Utility boilers
Commercial/industrial
boilers
Residential boilers
Municipal waste
combustors
Medical waste
incinerators
Sewage sludge
incinerators
Wood-fired boilers
Hazardous waste
combustors1
Crematories
Manufacturing
Chior-alkali production
Cement manufacturing
Battery production
Electrical apparatus
manufacturing
Instrument
manufacturing
Secondary mercury
production
Carbon black production
Primary lead smelting
Primary copper smelting*1
Lime manufacturing
Fluorescent lamp
recycling
Miscellaneous
Geothermal power plants
-
a Emissions in 1995 were estimated for hazardous waste incinerators and lightweight aggregate kilns;
however, these 1995 estimates were not used in any of the modeling analyses.
b Estimates were made for one source.only. This source ceased operations in February 1995.
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Table ES-3
Best Point Estimates of National Mercury Emission Rates by Category
Source of mercury3
Area sources
Lamp breakage
General lab use
Dental prep and use
Mobile sources
Paint use
Agricultural burning
Landfills
Point sources
Combustion sources
MWI/
MWCs
Utility boilers
Coal
Oil
Natural gas
Commercial/industrial boilers
Coal
Oil
Residential boilers
Coal
Oil
SSIs
Crematories
Wood-fired boilers"
Hazardous waste cotnbustors1
Manufacturing sources
Primary lead
Secondary Hg production
Color-alkali
Portland cement
Primary copper1
Lime manufacturing
Electrical apparatus
Instruments
Carbon black
Fluorescent lamp recycling
Batteries
Primary Hg production
Mercury compounds
Byproduct coke
Refineries
Miscellaneous sources
Geothermal power
Turf products
Pigments, oil, etc.
TOTAL
1990-1993
Mgtyf>-c
2J
1.4
0.7
0.7
d
e
d
d
2173
186.9
£8.8
50
46.5
(46.3)*
(0.23)
(0.002)
26.3
(20.7)
(5.5)
3.2
(0.5)
(2.7)
1.7
0.4
0.3
d
29.1
8.2
6.7
5.9
5.9
0.6
0.6
0.42
0.5
0.23
0.005
0.02
d
d
d
d
1.3
1.3
e
e
220.1
1990-1993
tons/yrb'c
3.1
1.5
0.8
0.8
d
e
d
d
239.4
205.9
64.7
55
-51.3
(51)
(0.25)
(0.002)
29
(22.8)
(6.0)
3.5
(0.6)
(3.0)
1.8
0.4
0.3
d
32
9.0
7.4
6.5
6.5
0.7
0.7
0.46
0.5
0.25
0.006
0.02
d
d
d.
d
1.4
1.4
e
e
242.5
% of Total
Inventory
1-J
0.6
0.3
0.3
d
e
d
d
98.7
84.9
26.7
22.7
21.2
(21.0)
(0.1)
(0.0)
12.0
(9.4)
(2.5)
1.4
0.2
(1-2)
0.7
0.2
0.1
d
13.2
3.7
3.1
2.7
2.7
0.3
0.3
0.2
0.2
0.1
0.002
0.0
d
d
d
d
0.6
0.6
e
e
100.0
' MWC = Municipal waste combustor, MWI = medical waste incinerator, SSI = sewage sludge incinerator.
b Numbers do not add exactly because of rounding.
° Where available, emissions estimates for 1995 are discussed in the text However, these 1995 estimates were not used in any of the modeling analyses.
* Insufficient information to estimate 1990 emissions.
' Mercury has been phased out of use.
f In the course of an MWI mlemalcing, with the receipt of new data, U.S. EPA expects to revise the mercury emission estimate for MWIs downward.
8 Parentheses denote subtotal within a larger point source category.
h Includes boilers only; does not include residential wood combustion (wood stoves).
1 In 1995 incinerators and lightweight aggregate kilns (not cement kilns) were estimated to emit 5.0 tons of mercury.
1 1990 emissions are estimated for only one source, which ceased operations in February 1995. The nationwide estimate for 1995 is 0.08 tons.
June 1996
ES-6
SAB REVIEW DRAFT
-------
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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 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 0.86 g
mercury per person per year. Based on data submitted during the 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).
The similarity between the U.S. inventory and other countries where coal is also the major source of
energy lends credibility to the nationwide estimate presented in this report for the U.S.
Trends in Mercury Emissions
It is difficult to predict with certainty the temporal trends hi 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 wherein mercury is a trace contaminant of the fuel.
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.
There has been a real success in the U.S. in the dramatic drop in mercury emissions from
manufacturing over the past decade. Current emissions of mercury from manufacturing sources are
generally low (with the exception of chlor-alkali plants using the mercury cell process). The emissions
of mercury are more likely to occur when the product is broken or discarded. Therefore, in terms of
emission trends, one would expect that if the future consumption of mercury remains consistent with
the 1993 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 two thirds between 1988 (1508 Mg) and 1993 (558 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
1993.
Secondary production of mercury (i.e., recovering mercury from waste products) has increased
significantly over the past few years. Of the 558 Mg of mercury used in industrial processes in 1993,
63 percent was provided by secondary mercury producers. This is a two-fold increase since 1991.
The number of secondary mercury producers is expected to increase as more facilities open to recover
June 1996 ES-12 SAB REVIEW DRAFT
-------
at
tL
OS
A
D
-------
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 identifiable sources of mercury emissions currently are municipal waste combustors
and medical waste incinerators. Emissions from these source categories are expected to decline
significantly 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 4.1.1 and 4.1.2 of this document, the
U.S. EPA has finalized rules for municipal waste combustors and proposed rules for medical waste
incinerators that will reduce mercury emissions from both of these source categories by about 90
percent 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.
After municipal solid waste and medical waste incinerators have been controlled, the largest
remaining identified source of mercury emissions will be 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 air quality
requirements under the Clean Air Act
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
dramatically over the past decade.
• Prior to 1995, municipal waste combustors and medical waste incinerators were the
largest identifiable source of mercury emissions to the atmosphere. Regulations
finalized for municipal waste combustors and proposed medical waste incinerators will
reduce emissions from these source categories by 90 percent.
• After emissions from municipal solid waste combustors and medical waste incinerators
have been reduced, combustion of fossil fuels, particularly coal, will be the largest
remaining source of mercury emissions to the atmosphere.
• 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.
• Anthropogenic sources in the United States emit approximately 220 Mg (243 tons) of
mercury annually into the atmosphere. This estimate is believed to be accurate to
June 1996 ES-14 SAB REVIEW DRAFT
-------
within 30 percent This estimate represents emissions calculated during the 1990-1993
timeframe. Emission estimates for 1995 are about 40 tons lower.
• In the United States, land areas east of the Rocky Mountains have the highest
concentration of emissions from anthropogenic sources in the U.S.
• The land 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 land 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.
• 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 bv 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.
June 1996 ES-15 SAB REVIEW DRAFT
-------
• Hie 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, landfills, 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, electric lamp breakage, iron and steel manufacturing and primary lead
smelting.
• Development and validation of a stack test protocol for speciated mercury emissions.
• More data on the efficacy of 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 Ihe 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 predict? the following:
• A significant (90 percent) decrease will occur in mercury emissions from municipal
waste combustors and medical waste incinerators when the regulations put forth by
U.S. EPA for these source categories are fully implemented.
June 1996 ES-16 SAB REVIEW DRAFT
-------
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.
June 1996 ES-17 SAB REVIEW DRAFT
-------
-------
1. INTRODUCTION
Section 112(n)(l)(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 a seven-volume Mercury Study Report to
Congress. The seven volumes are as follows:
I. Executive Summary
II. An Inventory of Anthropogenic Mercury Emissions in the United States
HI. An Assessment of Exposure from Anthropogenic Mercury Emissions in the United
States
IV. Health Effects of Mercury and Mercury Compounds
V. An Ecological Assessment of Anthropogenic Mercury Emissions in the United States
VI. Characterization of Human Health and Wildlife Risks from Anthropogenic Mercury
Emissions in the United States
VII. An Evaluation of Mercury Control Technologies and Costs
This volume (Volume n) 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 and state are
given for point sources.
This report updates the U.S. EPA document entitled National Emissions Inventory of Mercury
and Mercury Compounds: Interim Final Report (U.S. EPA, 1993d). The most significant changes to
the inventory are the utility boiler and municipal waste combustor source categories. Other source
categories (e.g., lamp breakage and chlor-alkali plants) were also updated. These changes are noted in
the text
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.
June 1996 1-1 SAB REVIEW DRAFT
-------
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.
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, 1993a). 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
June 1996 1-2 SAB REVIEW DRAFT
-------
Table 1-1
Sources of Anthropogenic Mercury Emissions Examined In This Inventory
Area
Electric lamp breakage
Paints use
Laboratory use
Dental preparations
Mobile sources*
Agricultural burning2
Landfills*
Sludge application*
Point
Combustion
Utility boilers
Commercial/industrial
boilers
Residential boilers
Municipal waste
combustors
Medical waste incinerators
Sewage sludge incinerators
Hazardous waste
combustors
Wood-fired boilers
Residential
woodstoves*
Crematories
Manufacturing
Chlor-alkali production
Lime manufacturing
Primary mercury production*
Mercury compounds
production
Battery production
Electrical apparatus
manufacturing
Carbon black production
Byproduct coke production*
Primary copper smelting
Cement manufacturing
Primary lead smelting
Petroleum refining*
Instrument
manufacturing
Secondary mercury
production
Zinc mining*
Fluorescent lamp recycling
Miscellaneous
Oil shale retorting
Mercury catalysts
Pigment production
Explosives manufacturing
Geothennal power plants
Turf products
a Potential anthropogenic sources of mercury for which emissions were not estimated.
June 1996
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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
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, 1993a). 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 populatioa 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 report 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 paniculate matter and acid gas controls for
utilities. Finally, Appendix C presents emission factors used to estimate emissions from utility boilers.
June 1996 1-4 SAB REVIEW DRAFT
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Table 1-2
Anthropogenic Mercury Sources With Sufficient
Data to Estimate National Emissions
Area
Electric lamp breakage
Laboratory use
Dental preparation
•
Point
Combustion
Utility boilers
Commercial/industrial
boilers
Residential boilers
Municipal waste \
combustors
Medical waste
incinerators
Sewage sludge
incinerators
Wood-fired boilers
Hazardous waste
combustors1
Crematories
Manufacturing
Chlor-alkali production
Cement manufacturing
Battery production
Electrical apparatus
manufacturing
Instrument
manufacturing
Secondary mercury
production
Carbon black production
Primary lead smelting
Primary copper smelting6
Lime manufacturing
Fluorescent lamp
recycling
Miscellaneous
Geothermal power plants
a Emissions in 199S were estimated for hazardous waste incinerators and lightweight aggregate kilns; however, these 199S
estimates were not used in any of the modeling analyses.
h Estimates were made for one source only. This source ceased operations in February 1995.
June 1996
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Table 1-3
Mercury Sources With Insufficient Information to Estimate National Emissions
Natural
Oceans and
other natural
waters
Vegetation
Volcanoes
Rocks
Soils
Wildfires
Anthropogenic
Area
Mobile sources
Paint use*
Agricultural
burning
Landfills %
Sludge
application
Point
Combustion
Residential
woodstoves
-
Manufacturing
Primary mercury
production1
Mercury compounds
production
Byproduct coke
production
Petroleum refining
Zinc mining
Miscellaneous
Oil shale retorting2
Mercury catalysts1
Pigment production2
Explosives
manufacturing3
Turf products*
* Mercury is no longer used in U.S. manufacture. However, this is not meant to imply that these previous activities are no
longer having an impact on the environment due to mercury's persistence in the environment
June 1996
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2. MERCURY TRENDS
Contemporary anthropogenic emissions of mercury are only one component of the global
mercury cycle. Releases from human activities today are adding to the mercury reservoirs that already
exist in land, water, and air, both naturally and as a result of previous human activities.
This chapter discusses the global mercury cycle and trends in mercury concentrations and
fluxes observed in the environment It also briefly reviews historical and recent patterns in mercury
supply and demand in the U.S. This discussion provides general background and context for the more
detailed discussion of anthropogenic sources in the chapters that follow.
2.1 Trends in the Atmospheric Mercury Burden
Given the present understanding of the global mercury cycle, the flux of mercury from the
atmosphere to land or "water at any one location is comprised of contributions from:
• The natural global cycle,
• The global cycle perturbed by human activities,
• Regional sources, and
• Local sources.
Understanding of the global mercury cycle (shown schematically in Figure 2-1) has improved
significantly with continuing study of source emissions, mercury fluxes to the earth's surface, and the
magnitude of mercury reservoirs that have accumulated in soils, watersheds and ocean waters.
Although considerable uncertainty still exists, it has become increasingly evident that anthropogenic
emissions of mercury to the air rival or exceed natural inputs. Recent estimates place the annual
amounts of mercury released into the air by human activities at between 50 and 75 percent of the total
yearly input to the atmosphere from all sources. Recycling of mercury at the earth's surface,
especially from the oceans, extends the influence and active lifetime of anthropogenic mercury releases
(Expert Panel, 1994).
A better understanding of the relative contribution of mercury from anthropogenic sources is
limited by substantial remaining uncertainties regarding the level of natural emissions as well as the
amount and original source of mercury that is re-emitted to the atmosphere from existing reservoirs.
Recent estimates indicate that of the approximately 200,000 tons of mercury emitted to the atmosphere
since 1890, about 95 percent resides in terrestrial soils, about 3 percent in the ocean surface waters,
and 2 percent in the atmosphere (Expert Panel, 1994). More study is needed before it is possible to
accurately differentiate natural fluxes from these reservoirs from re-emissions of mercury originally
released from anthropogenic sources. For instance, approximately one-third 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, a potentially large fraction of terrestrial and vegetative
emissions consists of recycled mercury from previously deposited anthropogenic and natural emissions
(Expert Panel, 1994).
June 1996 2-1 SAB REVIEW DRAFT
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Figure 2-1
The Global Mercury Cycle
Anthropogenic
Evasion
(Permitted
Anthropogenic
Natural)
Re-emitted
Anthropogenic
&
Natural
Local & Regional
Deposition
Global Terrestrial
Deposition
Global Marine
Deposition
Paniculate
Removal
Source: Adapted from Mason et al., 1994.
Comparisons of contemporary (within the last 15-20 years) measurements and historical
records indicate that the total global atmospheric mercury burden has increased since the beginning of
the industrialized period by a factor of between two and five (see Figure 2-2). For example, analysis
of sediments from Swedish lakes shows mercury concentrations in the upper layers that are two to five
times higher than those associated with pre-industrialized times. In Minnesota and Wisconsin, an
investigation of whole-lake mercury accumulation indicates that the annual deposition of atmospheric
mercury has increased by a factor of three to four since pre-industrial times. Similar increases have
been noted in other studies of lake and peat cores from this region, and results from remote lakes in
southeast Alaska also show an increase, though somewhat lower than found in the upper midwest U.S.
(Expert Panel, 1994).
While the overall trend in the global mercury burden since pre-industrial times appears to be
increasing, there is some evidence that mercury concentrations in the environment in certain locations
have been stable or decreasing over the past few decades. For example, preliminary results for eastern
red cedar growing near industrial sources (chlor-alkali, nuclear weapons production) show peak
mercury concentrations in wood formed in the 1950s and 1960s, with stable or decreasing
concentrations in the past decade (Expert Panel, 1994). Some results from peat cores and lake
sediment cores also suggest that peak mercury deposition occurred prior to 1970 (Benoit et al., 1994;
Swain et al., 1992; and Engstrom, 1994). Data collected over 25 years from many locations in the
United Kingdom on liver mercury concentrations in two raptor species and a fish-eating grey heron
indicate that peak concentrations occurred prior to 1970. The sharp decline in liver mercury
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Figure 2-2
Comparison of Current and Pre-Industrial
Mercury Budgets and Fluxes
Current Mercury
Budgets and
Fluxes
Anthropogenic
4.000
2,000
Air
5.000
98% Hg(g»MOu*) Hg(0)
2* Hg(particul»te)* 1
Local & Regional
Deposition
2,000
Global Terrestrial
Deposition
3.000
Global Marine
Deposition
2,000
Particulate
Removal
200
All Fluxes in 10'Kg/y
All Pools in 10* Kg
Pre-Industrial
Mercury Budgets
and Fluxes
Air
1,600
98% Hg(ga*aous) Hg(0)
2% Hg(particulate)
Global Terrestrial
Deposition
1,000
Global Marine
Deposition
600
Particulate
Removal
60
All Fluxes in 101 Kg/y
All Pools in 10* Kg
Source: Adapted from Mason et al., 1994.
June 1996
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concentrations in the early 1970s suggests that local sources, such as agricultural uses of fungicides,
may have led to elevated mercury levels two to three decades ago (Newton et al., 1993). Similar
trends have been noted for mercury levels in eggs of the common loon collected from New York and
New Hampshire (Mclntyre et al., 1993). This downward trend in mercury concentrations observed in
the environment over the last few decades generally tracks with mercury use and consumption patterns
over the same timefirame (discussed in the following section).
2.2 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 opera-
tions); (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 1993. The table
also shows the total industrial demand or consumption levels for that same period.
Figure 2-3 plots mercury supply and demand levels since 1955. Supplies associated with by-
product 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 2-3 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 50 percent since 1988.
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. Approximately 78 percent of the net U.S. exports of
mercury during the last five years has come from federal sales, with a steadily increasing portion of
the federal sales coming from the National Defense Stockpile managed by DLA. Federal sales
accounted for 97 percent of the U.S. demand in 1993 (Ross & Associates, 1994).
Most recently, there has been a sharp drop in Federal sales. 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).
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 hi
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 VII of this Report to Congress).
June 1996 2-4 SAB REVIEW DRAFT
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Table 2-1
U.S. Mercury: Supply, Demand, Imports, Exports
(Mg)
Category
Supply:
Mine production3
By-product production1*
Industrial recovery
DLA sales
DOE sales
Subtotal: federal sales
Imports
Total supply
Federal sales as % of total
supply
Demand:
Federal sales as % of domestic
demand
Imports:
Exports:
Exports minus imports:
1988
379
W°
278
52
214
266
329
1,252
21.2%
1,503
17.6%
329
N/Ad
N/A
1989
414
W
137
170
180
350
131
1,032
33.9%
1,212
28.9%
131
221
90
1990
448
114
108
52
193
245
15
930
26.3%
720
34%
15
311
296
1991
0
58
165
103
215
318
56
597
53.3%
554
57.4%
56
786
730
1992
0
64
176
267
103
370
92
702
52.7%
621
59.6%
92
977
. 885
1993
0
W
350
543
0 '
543
40
933
58.1%
558
97.3%
40
389
349
Source: U.S. Bureau of Mines, 1994.
a Mercury production from McDennitt mine; closed November 1990.
b Mercury by-product from nine gold mining firms.
c Withheld for proprietary reasons.
d Not available.
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3. ANTHROPOGENIC AREA SOURCES OF MERCURY
EMISSIONS
Area sources account for approximately 1.3 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 high-
pressure sodium lamps. 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 mem 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 hi the lamp and the size of the paniculate forms of mercury in
the lamp powder. Approximately 643 Mg of mercury were discarded hi U.S. municipal solid waste
(MSW) in 1989. The amount of mercury entering the MSW system from the disposal of used
mercury-containing 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).
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, one manufacturer
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 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.
June 1996 3-1 SAB REVIEW DRAFT
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Table 3-2
Mercury Content of Fluorescent Bulbs'
Year
1970-1984
1985-1989
1990
1992
1995
Average Mercury Content (mg) per Bulb
NEMA CWF
75
482 55
41.6
40
22.8
3M
15-30
* Cole et al., 1992; National Electrical Manufacturers Association, 1992; Tanner, 1992; National Electrical
Manufacturers Association, 1995.
Table 3-3
Mercury (HID) Lamp Production -1970 to 1989*
Year
1970
1971
°1972
1973
1974
1975
1976
1977
1978
1979
1980
1981
Quantityb (1000 bulbs)
6,841
7,684
8,420
9,349
9,158
8,737
10,383
10,853
12,175
13,532
30,187
21,397
Year
1982
1983
1984
1985
1986
1987
1988
1989
Quantity1" (1000 bulbs)
20,891
22,146
25,636
25,529
22,206
28,143 -
24,479
28,090
* Cole et al., 1992; U.S. EPA, 1992a.
b Production rate = Domestic shipments - Exports + Imports.
June 1996
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Table 3-4
Mercury Content of HID Lamps*
Manufacturer
Phillips
Sylvania
Type
250 watt HE)
400 watt HE)
1000 watt HE)
250 watt HE)
400 watt HE)
1000 watt HE)
Mercury Content (ing)
45
60
70
46
75
75
a Cole et aL, 1992; U.S. EPA, 1992a.
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).
The fate of used lamps is tied to the disposal of MSW. The three primary options for MSW
disposal are landfilling, combustion and recycling. Landfilling 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. Ninety-eight
percent of used lamps are managed as MSW under Subtitle D (the solid, non-hazardous 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). 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. Insufficient
data exist to estimate mercury emissions from landfills.
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 relamping 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.
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For transportation in a garbage truck, it was assumed that all lamps are broken in the truck and
that the mercury present in the 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).
The total mercury emissions from mercury lamp breakage, including emissions from transport
and storage, are estimated to be 1.4 Mg/yr (1.5 tons/yr). This estimate may have a high degree of
error associated with it due to the number of assumptions made and the lack of test data. This
estimate represents a change from U.S. EPA's earlier estimate of 8 Mg (8.8 tons/year) (U.S. EPA,
1993d). The current estimate is based on more plausible assumptions than the earlier analysis which
assumed a large fraction of the mercury was emitted to the atmosphere upon lamp breakage. These
earlier estimates were though to be high considering that only a small percentage of mercury in the
lamp is in the vapor state.
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 landfilled within a couple of days after their disposal and are covered with 0.5 to
1 foot of soil at that time. Simulating this landfilling 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 1992, an
estimated 0.7 Mg (0.8 ton) 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. As with most industries using mercury, there was a decline in
mercury consumption in general laboratory use, with consumption dropping from 32 Mg (35 tons) in
June 1996 3-6 SAB REVIEW DRAFT
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1990 to 18 Mg (20 tons) in 1992 (Bureau of Mines, 1992). 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.
3J 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 1992, an estimated 0.7 Mg (0.8 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 37 Mg (41 tons) and includes mercury used in
all dental equipment and supplies, not just the amount used in dental preparation and use (Bureau of
Mines, 1992). Mercury emissions not accounted for in dental preparation and use are most likely
accounted for in the emission estimates for municipal waste combustors anNl crematories. Mercury
discharges from dental offices to publicly owned sewage treatment faculties are also known to occur
but are not addressed in this report.
3.4 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 10'3 milligram per
kilometer (mg/km) (4.6 x 10"9 pound per mile Qb/mile]) traveled for tail-pipe emissions from motor
vehicles (Pierson and Brachaczek, 1983). These data were for paniculate mercury emissions derived
from neutron activation analysis of paniculate filters. The population of vehicles studied was
81.9 percent gasoline-powered passenger cars, 2.4 percent gasolinerpowered 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
conveners 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
non-catalyst 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 10'8 Ib/mile) HFET 0.019 mg/km (6.7 x lO'* Ib/mile) and NYCC
0.15 mg/km (53.2 x 10'8 Ib/mile) (Warner-Selph and Lapp, 1993). These minimum detection limits
are more than ten times 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
June 1996 3-7 SAB REVIEW DRAFT
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estimated in this study. The potential for mercury emissions from other types of mobile sources,
including ships, were not assessed in this study.
3.5 Paint Use
Four mercury compounds — phenylmercuric acetate, 3-(chloromethoxy) propylmercuric acetate,
di(phenylmercury) dodecenylsuccinate, and phenylmercuric oleate - have been registered as biocides
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
interiof paint in 1990 and in exterior paint in 1991. The pJunt 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 hi 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 mat 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 hi paint was high, estimated that 66 percent of the mercury
used hi 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.
3.6 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 hi 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
land-clearing 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.
June 1996 3-8 SAB REVIEW DRAFT
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3.7 Other Area Sources
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 landfilling stops
and waste materials are covered with soil. Although mercury emissions have been estimated for
fluorescent lamps discarded in landfills (see Section 3.1), insufficient information was available to
estimate mercury emissions from landfills more generally.
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.
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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 99 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.
Table 4-1
Best Point Estimates of Annual Mercury Emissions from Combustion, Manufacturing and
Miscellaneous Point Sources: 1990-1993
Source
Combustion
Manufacturing
Miscellaneous
Total Point Source Emissions
Emissions
Mg/yr
186.9
29.1
1.3
217.3
Tons/yr
205.9
32.0
1.4
239.4
4.1 Combustion Sources
Combustion sources include medical waste incinerators, municipal waste combustors, fossil
fuel-fired boilers, residential boilers, wood combustion, sewage sludge incinerators and crematories.
Mercury emissions from these sources (excluding wood-fired residential heaters) account for an
estimated 186.9 Mg/yr (205.9 tons/yr) or 85 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 locations 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) medical waste incinerators (MWIs), (2) municipal waste combustors (MWCs), (3) utility boilers,
(4) commercial/industrial boilers, (5) residential boilers, (6) sewage sludge incinerators (SSIs), (7)
crematories, (8) wood combustors, and (9) hazardous waste combustors. For each of these source
types, processes and control measures currently in place are discussed, along with emission estimates
June 1996
4-1
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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.
4.1.1 Medical Waste Incinerators
Medical waste incinerators (MWIs) are small- incineration units that charge from 0.9 Mg/d
(1 ton/d) to 55 Mg/d (60 tons/d) 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).
o
The estimated annual uncontrolled mercury emissions from MWIs are currently 58.8 Mg/yr
(64.7 tons/yr). In the course of developing NSPS and emission guidelines for MWIs, with the receipt
of new data, U.S. EPA expects to revise this estimate of current emissions downward. In addition, the
recently proposed NSPS and emission guidelines for MWIs would decrease national mercury emissions
from MWIs by almost 85 percent, to an estimated level of 3.7 Mg/yr (4.1 tons/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
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 106 Mg/yr (0.268 x 106 tons/yr) of pathological waste
and 1.431 x 106 Mg/yr (1.574 x 106 tons/yr) of general medical waste are processed annually in the
United 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.
As of 1985, there were a reported 6,872 known hospitals in the United States. Manufacturers
estimate that at least 90 percent of these hospitals have an incinerator on-site, if only a small retort-
type for burning pathological wastes (Epner, 1992). The number of larger, controlled-air MWIs is not
June 1996 4-2 SAB REVIEW DRAFT
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New Source Performance Standards and
Emission Guidelines for MWIs
On February 27, 1995, EPA proposed NSPS for new MWIs and emission guidelines for existing
MWIs (60 FR 10654). The standards would apply to all new and existing MWIs that commence- construction
or modification after February 27, 1995, although sources combusting only pathological wastes would be
subject to only certain reporting and recordkeeping provisions. Overall, the NSPS and emission guidelines
would 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).
The proposed NSPS would reduce emissions from MWIs by establishing standards that limit
emissions from new MWIs. For mercury, the standard would be 0.47 mg/dscm or an 85 percent reduction in
mercury emissions (based on dry injection/fabric filter with an activated carbon injection system). The
proposed NSPS also would 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 recordkeeping requirements.
The proposed emission guidelines would reduce emissions from MWIs by requiring States to develop
regulations that limit emissions from existing MWIs. The mercury emission limit proposed in the guidelines is
the same as that proposed in the NSPS, i.e., 0.47 mg/dscm or an 85 percent reduction in mercury emissions
(based on dry injection/fabric filter with an activated carbon injection system). Consistent with the NSPS, the
emission guidelines also would require training and qualification of MWI operators, specify testing and
monitoring requirements, and establish reporting and recordkeeping 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, provided the owner or
operator submits information on measurable and enforceable incremental steps of progress that will be taken to
comply with the State plan.
known. An estimated 1,200 systems have been installed at U.S. hospitals in the past 20 years (Epner,
1992). About 3,700 MWIs currently operate throughout the country; geographic distribution is
relatively even (see Table A-ll, Appendix A). Of these 3,700 units, about 3,000 are hospital
incinerators, about 150 are commercial units, and the remaining units are distributed among nursing
homes, laboratories and other miscellaneous facilities (U.S. EPA, 1994a).
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 58.8 Mg/year (64.7
tons/year). The reader should note that the proposed 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 4.1.7.
The primary functions of MWIs are to render the waste biologically innocuous and to reduce
the volume and mass of solids that must be landfilled 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
June 1996 4-4 SAB REVIEW DRAFT
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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
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 sorbent injection/fabric filtration systems can achieve improved mercury control
by adding activated carbon to the sorbent material (U.S. EPA, 1993a). (Activated carbon injection for
mercury control at MWIs is discussed in Volume YE of this Report to Congress.)
The estimated mercury emission factors for MWIs were determined by analyzing test data
from several MWI facilities tested by U.S. EPA. The uncontrolled emission data collected at eight
facilities show wide variability, with mercury emission factors ranging from 0.043 to 317 g/Mg of
waste charged (8.6 x 10"4 to 6.3 x 10'1 Ib/ton) (U.S. EPA, 1993a). These data represent a variety of
waste types (mixed medical waste, red bag [infectious] waste only and pathological waste) and
incinerator types (continuous and intermittent units with varied operating practices). Because
emissions are strongly related to waste characteristics, separate uncontrolled emission factors were
developed for the different waste types. The average uncontrolled mercury emission factor for mixed
medical waste was calculated to be 20 g/Mg of waste charged (4.0 x 10*2 Ib/ton). The average
uncontrolled mercury emission factor for red bag waste was calculated to be 16 g/Mg of waste charged
(3.2 x 10"3 Ib/ton). The average uncontrolled mercury emission factor for pathological waste was
calculated to be 5 x 10*3 g/Mg waste charged (1 x IQ'3 Ib/ton) (U.S. EPA, 1993a).
Mercury emission estimates for the MWI source category were calculated using model units.
Model units are representative facilities that are developed to represent an industry when information
from each plant is unavailable. Seven model units were developed to represent the variety of MWIs in
the United States (U.S. EPA, 1994b). The models differed by type of MWI operation (continuous,
batch and intermittent), waste feed rate and operating hours, among other parameters. All seven model
units represent uncontrolled MWIs. The aggregate uncontrolled mercury emissions from the MWI
industry were estimated by multiplying the number of MWIs assigned to each model, the emission
factors, feed rates and operating hours. Table A-12 in Appendix A presents the characteristics of each
model MWI, the number of MWIs in the U.S. that were assigned to each model and the estimated
mercury emissions from them. The total uncontrolled mercury emissions were estimated to be
58.8 Mg/yr (64.7 tons/yr). As mentioned above, with the receipt of more recent data, the U.S. EPA
expects to revise this estimate downward.
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 bum combustors, modular combustors and refuse-derived fuel-fired (RDF) combustors. A fourth
type, fiuidized-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
June 1996 4-5 SAB REVIEW DRAFT
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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 and their capacities are
presented in Table A-9, Appendix A (Fenn and Nebel, 1992). This distribution reflects MWC's that
were operational hi 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 hi this size range that have closed. Table A-9 in Appendix A, as
well as the map shown in Figure 4-1, reflects the 1995 MWC population.
Figure 4-1
Municipal Waste Combustor Facilities
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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 Gess than 150 to 200°C [300 to 400°F]), the presence of an
effective mercury sorbent and a method to collect the sorbent (Nebel and White, 1992). In general,
carbon present hi the fly ash enhances mercury sorption onto PM, which can then be captured in the
PM control device. Most modern 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 NajS 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 efficiencies 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 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
will require MWCs in these states to retrofit with activated carbon injection by the end of 1995.
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-10, Appendix A. Estimated mercury emissions were determined based on the tonnage of the
waste being combusted (Table A-9 in Appendix A) 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-10, 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.
Industry estimates of mercury emissions from this source category for 1990 are 40 Mg/year
(44 tons/year) (Kiser, 1991). The difference between the assumptions that U.S. EPA used and the
assumptions that industry used are described in Table A-9, Appendix A. The primary differences in
June 1996 4-7 SAB REVIEW DRAFT
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the assumptions are related to the average mercury emission rate and the extent of mercury reduction
that can be attributed to an ESP.
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).
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
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 being
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 26.3 Mg/yr (29 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 VII of this
Mercury Report to Congress summarizes the legislative, regulatory and other programs of several
states that influence mercury use and disposal.
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
June 1996 4-8 SAB REVIEW DRAFT
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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 recent study identified and reported a number of specific sources of mercury in MSW, as
summarized in Tables 4-3 and 4-4. The data, from Table 4-4 are shown graphically for the year 1989
in Figure 4-2. These figures show that in 1989 household batteries were the largest contributing
source of mercury to MSW. 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 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).
Mercuric oxide batteries include cylinder-shaped batteries (such as those used in hospital
applications) and button-shaped batteries (such as those used hi hearing aids, electronic watches,
calculators, etc.). Larger mercuric oxide batteries are used hi 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 hi MSW on a
percentage basis (Solid Waste Association of North America, 1993), the total, tonnage of mercury
discarded hi such batteries is expected to decline in the future due to the increase in use of alkaline
and zinc air batteries for these applications. Presently, mercuric oxide batteries are essentially the
remaining source of mercury in MSW from batteries.
June 1996 4-9 SAB REVIEW DRAFT
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Table 4-3
Estimated Discards of Mercury in Products in Municipal Solid Waste (in tons)3
Products
Batteries
Alkaline
Mercuric oxide
Others
Subtotal Batteries
Electric Lighting
Fluorescent Lamps
High Intensity
Lamps
Subtotal Lighting
Paint Residues
Fever Thermometers
Thermostats
Pigments
Dental Uses
Special Paper Coating
Mercury Light Switches
Film Pack Batteries
Total Discards
In Tonsb>c
1970
4.1
301.9
4.8
310.8
18.9
0.2
19.1
30.2
12.2
53
323
93
0.1
0.4
2.1
421.8
1975
38.4
287.8
4.7
330.9
21.5
03
21.8
373
23.2
6.8
27.5
9.7
0.6
0.4
2.3
460.5
1980
158.2
266.8
4.5
429.5
23.2
1.1
243
26.7
25.7
7.0
23.0
7.1
1.2
0.4
2.6
547.5
1985
3523
235.2
4.5
592.0
27.9
0.7
28.6
31.4
32 .5'
9.5
25.2
6.2
1.8
0.4
2.8
730.4
1989
419.4
196.6
5.2
621.2
26.0
0.8
26.7
18.2
16.3
11.2
10.0
4.0
1.0
0.4
0.0
709.0
1995
41.6*
1313*
3.5*
176.6*
e>
14.7d
1.0
15.7
2.3
16.9
8.1
3.0
2.9
0.0
1.9
0.0
227.6
2000
0.0
98.5*
0.0
98.5*
11. 6d
1.2
12.6
0.5
16.8
103
1.5
23
0.0
1.9
0.0
144.6
11 U.S. EPA, 1992a (except for fluorescent lamps estimates).
Discards before recovery.
0 One ton equals 2000 pounds.
d The estimated contribution of mercury from fluorescent lamps disposal to MSW was calculated based on industry
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: 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. Since the referenced report was released (U.S. EPA, 1992a), 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 is pending. The battery industry has eliminated mercury as an
intentional additive in alkaline batteries, except in button cells.
June 1996
4-10
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Table 4-4
Estimated Discards of Mercury in Products in Municipal Solid Waste (in percent) a
Products
Batteries0
Alkaline
Mercuric oxide
Others
Subtotal Batteries
Electric Lighting
Fluorescent Lamps
High Intensity Lamps
Subtotal Lighting
Paint Residues
Fever Thermometers
Thermostats
Pigments
Dental Uses
Special Paper Coating
Mercury Light Switches
Film Pack Batteries
Total Discards
1970
1.0
71.6
1.1
73.7
4.5
0.0
4.5
7.2
2.9
1.3
7.7
2.2
0.0
0.1
0.5
100.0
1975
8.3
62.5
1.0
71.9
4.7
0.1
4.7
8.1
5.0
1.5
6.0
2.1
0.1
0.1
0.5
100.0
In Percent
1980
28.9
48.7
0.8
78.4
4.2
0.2
4.4
4.9
4.7
1.3
4.2
1.3
0.2
0.1
0.5
100.0
of Total
1985
482
322
0.6
81.1
3.8
0.1
3.9
4.3
4.4
1.3
3.5
0.8
0.2
0.1
0.4
100.0
Discards*
1989
59.2
27.7
0.7
87.6
a
3.7
0.1
3.8
2.6
2.3
1.6
1.4
0.6
0.1
0.1
0.0
100.0
1995
17.0*
53.6*
1.4*
72.0*
13.3
0.4
13.7
0.9
6.9
3.3
1.2
1.2
0.0
0.8
0.0
100.0
2000
0.0
68.0*
0.0
68.0*
8.0
0.8
8.7
0.4
11.6
7.1
1.0
1.6
0.0
1.3
0.0
100.0
4 U.S. EPA, 1992a.
b Discards before recovery.
0 Data were based on the assumption that batteries are discarded 2 years after purchase.
Details may not add to totals due to rounding.
* NOTE: The estimates for the years 1995 and 2000 do not reflect recent state, Federal 01 battery manufacturers'
efforts to reduce the mercury content of batteries. Since the referenced report was released! (U.S. EPA, 1992a), 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 is pending. The battery industry has eliminated mercury as an
intentional additive in alkaline batteries, except in button cells.
June 1996
4-11
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Figure 4-2
Discards of Mercury in Municipal Solid Waste, 1989
All Others 1%
^^___ji_ Pigments 1.4%
\Batteries 87.3% "^^^MTs^Thermostats 1.6%
^^™* \Thermometers 2.3%
Paint Residues 2.6%
Lighting 3.8%
Total mercury discards = 709 tons
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).
Table 4-5 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.
Recent efforts to reduce the mercury content of batteries have included either passed or
pending legislation in (at the time of this writing) at least 18 states. As just one example, Minnesota
has passed a law requiring manufacturers to sell alkaline batteries containing no more than 0.025
percent mercury by weight and, by January 1, 1996, prohibiting manufacturers from selling alkaline
batteries containing any mercury. Efforts at the Federal level include pending legislation that would
eliminate the use of mercury in all types of batteries used in both industrial and household activities
by January 1, 1997. The battery industry, as mentioned previously, also has eliminated mercury as an
intentional additive in alkaline batteries, except in button cells. Several communities have established
programs to recover household batteries in order to prevent the disposal of batteries in landfills or
incinerators. At the time of this writing, three U.S. firms had the facilities necessary to recycle the
mercury in batteries. U.S. EPA estimates that approximately 5 percent of the mercury in batteries will
be recovered through recycling in 1995 and approximately 20 percent will be recovered by 2000.
June 1996 4-12 SAB REVIEW DRAFT
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Table 4-5
Estimated Discards of Mercury in Batteries*
InTonsb
Alkaline
4.1
38.4
158.2
352.3
443.6
390.5
41.6*
0.0
Mercuric Oxide
301.9
287.8
266.8
2352
182.5 %
172.0
131.5*
98.5*
Silver Oxide
0.1
0.2
0.3
0.5
1.1
1.1
0.7
0.0
Zinc Air
0.0
0.2
0.3
0.7
2.4
2.9
2.0
0.0
Year Discarded
1970
1975
1980
1985
&90
1991
1995
2000
* U.S. EPA, 1992a.
b One ton equals 2000 pounds.
* 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. Since the referenced report was released (U.S. EPA, 1992a), 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 is pending. The battery industry has eliminated mercury as
intentional additive in alkaline batteries, except in button cells.
Chapter 4 of Volume VII of this Mercury Study Report to Congress provides! additional discussion of
various laws and programs for controlling battery production, use and disposal.
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) hi 1989. All lighting sources were estimated to contribute about 27 tons of mercury
in the same year. Figure 4-3 illustrates the estimated historical discards of electric lighting sources.
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 3.2 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
hi 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.
June 1996 4-13 SAB REVIEW DRAFT
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Figure 4-3
Estimated Discards of Mercury in Electric Lighting in Municipal Solid Waste
tons
45 y
40--
35-.
30--
25--
15-
10-
5-
0-
•
•
•
1970
1975
1980
1985
1990
Paint Residues
A
Mercury is no longer used in paint manufacture; however, paint cans with traces of mercury
are still discarded. It was estimated that about 18 tons of mercury were discarded in paint residues in
1989. Mercury from paint residues is expected to decline significantly due to U.S. EPA's recent ban
on mercury use in interior and exterior paints. Table 4-6 presents estimated mercury discards from
paint residues from 1970 to 2000.
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). Tables 4-3 and 4-4 illustrates the estimated discards of mercury from thermometers in MSW
from 1970 to 2000.
June 1996
4-14
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Table 4-6
Estimated Discards of Mercury in Paint Residues'*
Year
1970
1975
1980
1985
1988
1990
1995
2000
Total Discards in Residues (In Tons)
30.2
37.3
26.7
31.4
23.1
17.5
2.3
0.5
* U.S. EPA, 1992a.
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 4-7. Federal legislation finalized in 1995 will encourage the recycling
of thermostats rather than their disposal. Recyclings efforts are discussed in section 4.2.7.1 of this
Volume. As a result of recycling programs, mercury discards from thermostats are expected to
decline.
June 1996 4-15 SAB REVIEW DRAFT
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Table 4-7
Estimated Discards of Mercury in Thermostats3
Year
1970
1975
1980
1985
1988*
1989
1995
2000
Total Mercury (In Tons)
5.3
6.8
7.0
9.5
10.7
11.2
8.1
10.3
• U.S. EPA, 1992a.
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 4-4, estimated discards of mercury in MSW pigments have generally been trending downwards
since 1970 (U.S. EPA, 1992a).
June 1996 4-16 SAB REVIEW DRAFT
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Figure 4-4
Estimated Discards of Mercury in Pigments in Municipal Solid Waste
torn
60 T
50-
40- .
30.
20 =
10-
•4-
1970 1975 1980 1985 1990 1995 2000
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.
•
In the production of paper, mercury compounds were formerly used as slimicides to prevent
the growth of green slime on the manufacturing equipment. Mercury compounds also were used to
prevent the growth of mold and bacteria on pulp during storage, but this practice has been
discontinued (U.S. EPA, 1992a).
4.1.3 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).
Figure 4-5 shows the locations of operating utility boilers across the United States.
In 1990, utility boilers consumed fossil fuel at an annual level of 21 x 1012 megajoules (MJ)
(20 x 1015 British thermal units [Btu]). About 80 percent of this total energy consumption resulted
from coal combustion, 6 percent from oil and petroleum fuels and 14 percent from natural gas
consumption (U.S. Department of Energy, 1992). In terms of coal usage, the majority of total
nationwide coal combustion (about 84 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
June 1996 4-17 SAB REVIEW DRAFT
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m o
4 .3
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oi
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1
03
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fuel used by utility boilers is fuel oil derived from crude petroleum. Fuel oils are classified as either
distillate or residual.
4.1.3.1 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/1993a).
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
suspension-fired 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).
4.1,3.2 Effectiveness of Paniculate Matter and Acid Gas Air Pollution Controls for Mercury
Although small quantities of mercury may be emitted as fugitive paniculate 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 1100°C (2000°F), 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 4-6. 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 16.2 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; and Western, 1993b). 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
June 1996 • 4-19 SAB REVIEW DRAFT
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35
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2f
CO
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2
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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 SO2 emissions). Most utility boilers have an ESP or a FF before the wet FGD units to
collect the majority of PM. FGD units have a median mercury removal efficiency of about
17.3 percent, with a range from 0 percent to 59 percent removal (Interpoll Laboratories, 1991;
InterpoU Laboratories, 1990a; Radian, 1993a; Radian, 1993b; 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 paniculate 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).
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 man or equal to 1 urn in size. These devices are used as gross paniculate
removal devices before ESPs or as APCDs on oil-fired units. Venturi scrubbers can be effective for
paniculate 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.
4.1.3.3 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
boiler-specific basis) were obtained from the Utility Data Institute (UDI)/Edison Electric Institute (EEI)
Power Statistics database (1991 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
June 1996 4-21 SAB REVIEW DRAFT
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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 Ib/gal was chosen to represent all
residual oils. Trace element analysis of natural gas was performed for only two available emissions
tests; these concentrations were averaged. The calculated mercury concentration hi 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
June 1996 4-22 SAB REVIEW DRAFT
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associated mercury - may be removed by coal cleaning processes. Since approximately 77 percent of
the eastern and midwestera 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.
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 51 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.2
Figures 4-7 and 4-8 illustrate how mercury emission estimates were calculated for coal-fired boilers
and for oil- or natural gas-fired boilers. As displayed in Table 4-9, national estimates of mercury
emissions from utility boilers are 51.3 tons per year, of which 51 tons are attributed to coal-fired units,
0.25 tons are attributed to oil-fired units, and 0.002 tons are attributed to natural gas-fired units.
4.1.4 Commercial/Industrial Boilers
Commercial/industrial boilers are large boilers found in businesses and industrial plants
throughout the United States. These boilers may use coal, oil, or natural gas as fuels. As with utility
boilers, mercury vaporizes during combustion and appears as a trace contaminant in the gas exhaust
stream.
2 Limestone is used in circulating fluidized bed (CFB) boilers to control sulfur dioxide emissions. The EPA
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.
June 1996 4-23 SAB REVIEW DRAFT
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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 <1afa 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 Stale 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 <*f>ail subbituminous Wyoming, and bituminous Illinois coal shipments.
The mercury concentrations in the raw coal, the dean coal, and the percent reduction achieved by
cleaning are shown in Table 4-8. 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 4-8, the mercury redactions 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 talcing the
arithmetic average of the removal data listed in Table 4-8. 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 bum bituminous coal from states east of the
Mississippi River. Note that mis reduction was assumed for all such boilers, even though data indicate
that only 77 percent of the eastern and midwestem 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.
June 1996 4-24 SAB REVIEW DRAFT
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Figure 4-7
Mercury Emissions from Oil- and Natural-Gas Fired Boilers
OIL NATURAL GAS
PLANT
CONFIGURATION
INFORMATION
PLANT
CONFIGURATION
INFORMATION
Used fuel oil #6
(residual) for all oil
types
Trace elements in
oil taken from
plant testing
Used a density of
8.2 Ib/gal for feed
rate calculation
Trace elements in
gas taken from
plant testing
(only 2 sets of data)
Apply boiler trace element
emission factors (EMFs)
What type of paniculate
matter (PM) control?
Apply PM trace element
emission factors
What type of SO, control?
Apply SO, trace element
emission factors
June 1996
I Kg/yr mercury out of stack I
4-25
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Figure 4-8
Mercury Emissions from Coal-Fired Boilers
USGS average coal
mercury concentration
specific to State of
coal origin
PLANT
CONFIGURATION
INFORMATION
Identify State of coal
origin from UDI
USGS average coal
mercury concentration
specific to State of
coal origin
Apply coal
cleaning factor,
if applicable
Multiply mercury content
of coal by unit annual feed
rate form UOI data base
What type of boiler?
Apply boiler trace element
emission factors (EMFs)
What type of participate
matter (PM) control?
Apply PM trace element
emission factors
What type of SO, control?
Apply SO, trace element
emission factors
Kg/yr mercury out of stack
June 1996
4-25a
SAB REVIEW DRAFT
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Table 4-8
Comparison of Mercury Concentrations in Raw and Cleaned Coal
Refei
Seam
Central Appalachian Coal Sample A
Central Appalachian Coal Sample B
11 #6
Pittsburgh A
Pittsburgh B
Pittsburgh C
Pittsburgh D
Pittsburgh E
Pittsburgh
Upper Freeport
Lower Kittanning
Sewickley
Pittsburgh
Pittsburgh
11 #6
KY #9 and 14
Pratt/Utley
Pratt
UUey
Pratt
Upper Freeport
Upper Freeport
11 2,3,5
11 2,3,5
Ky#ll
ISGS
Minimum
Maximum
Average
State
IL
PA
PA
PA
PA
PA
PA
PA
PA
PA.
PA
PA
IL
KY
AL
AL
AL
AL
PA
PA
IL
IL
KY
IL
Raw Coal
Mercury (ppm)
0.09
0.12
0.14
0.15
0.14
0.14
0.1
0.1
0.1
0.03
0.44
0.18
0.13
0.13
0.12
0.16
0.28
0.29
0.34
0.34
0.7
0.7
0.24
0.24
0.15
0.2
Cleaned Coal
Mercury (ppm)
0.1
0.11
0.08
0.11
0.09
0.13
0.12
0.08
0.08
0.09
0.34
0.18
0.11
0.12
0.13
0.14
0.22
0.28
0.27
0.24
0.25
0.28
0.2
0.14
0.12
0.09
Percent
Removal
-11.11
8.33
42.86
26.67
35.71
7.14
-20.00
20.00
20.00
-200.00
22.73
0.00
15.38
7.69
-8.33
12.50
21.43
3.45
20.59
29.41
64.29
60.00
16.67
41.67
20.00
55
-200.00
64.29
21.21
•ence: Akers et al., 1993 for every seam but ISGS; Demir et al., 1993 for ISGS.
June 1996
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Table 4-9
Best Point Estimate of Mercury Emissions from Utility Boilers: 1990
Fuel Type
Coal
Oil
'Natural Gas
Total
Emission Rate
Mg/Yr
46.3
0.23
0.002
46.5
Tons/Yr
51
0.25
0.002 \
51.3
Comments
The industry (Electric Power Research
Institute) estimate for coal-fired units is 44
tons/year.
%
Mercury emissions from commercial/industrial boilers, estimated at 26.3 Mg/yr (29 tons/yr),
are directly related to the amount of fuel used in the combustion process (U.S. EPA, 1993a). Mercury
emissions from natural gas combustion could not be estimated because a reliable emission factor does
not exist (U.S. EPA, 1993a). Commercial/industrial boilers consume energy at an annual rate of
25 x 1012 MJ/yr (23 x 1015 Btu). About 12 percent of this energy consumption results from coal
combustion, 39 percent from oil and petroleum fuel combustion, and 48 percent from natural gas
combustion (U.S. Department of Energy, 1992). Estimates of coal and oil consumption from these
boilers on a per-State basis are presented in Table A-2, Appendix A.
Because there is no evidence to show that mercury emissions are affected by boiler type, this
section presents only a brief discussion of commercial/industrial boiler types and combustion
techniques. More 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, 1988a; U.S. EPA, 1993a).
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 mose used in the utility sector. Larger coal-fired
industrial boilers are suspension-fired systems like those used hi 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 faculties, also have ash-handling systems.
Mercury emission factors for coal combustion hi commercial/hidustrial 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, 1993a). 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/1015 J (16 lb/1012 Btu) was used for bituminous coal combustion, and 7.6 kg/1015 J
(18 lb/1012 Btu) was used for anthracite coal combustion. Estimates of mercury emissions on a per-
June 1996
4-27
SAB REVIEW DRAFT
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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 20.7 Mg/yr (22.8 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
per-state basis using an emission factor of 2.9 kg/1015 J (6.8 lb/1012 Btu) for residual oil and
3.0 kg/1015 J (7.2 lb/1012 Btu) for distillate oil and the oil consumption estimates for States given in
Table A-2, 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 5 Mg/yr (6 tons/yr).
4.1.5 Residential Boilers
Residential boilers are relatively small boilers used hi 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
oil-fired residential boilers and the emissions appear as a trace contaminant hi the exhaust gas.
The estimated annual mercury emissions from residential boilers, 3.2 Mg/yr (3.5 tons/yr), are
related to the amount of fuel used hi the combustion process. Estimates of coal and oil consumption
from these boilers on a per-state basis are presented in Table A-4, Appendix A. Residential boilers
consume energy at an annual rate of 6.2 x 1012 MJ/yr (5.8 x 1015 Btu/yr). About 1 percent of this
energy consumption results from coal combustion, 22 percent from oil and petroleum fuel combustion
and 77 percent from natural gas combustion (U.S. Department of Energy, 1992).
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).
Estimated mercury emission factors for coal combustion in residential boilers are the same as
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/1015 J
(16 lb/1012 Btu) was used for bituminous coal combustion, and 7.6 kg/1015 J (18 lb/1012 Btu) was
used for 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 hi Table A-5, Appendix A. These calculated emission values are
presented in Table A-6, Appendix A. hi 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 0.5 Mg/yr (0.5 tons/yr).
The estimated mercury emissions for oil combustion were estimated by using an emission
factor of 2.9 kg/1015 J (6.8 lb/1012 Btu) for residual oil and 3.0 kg/1015 J (7.2 lb/1012 Btu) for
distillate oil and 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 2.7 Mg/yr (3.0 tons/yr).
June 1996 4-28 • SAB REVIEW DRAFT
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4.1.6 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 mercury in sewage conies 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 landfilled or incinerated. Only a small percentage of U.S. cities use sewage sludge
incinerators. The estimated annual mercury emissions from SSIs account for 1.65 Mg/yr
(1.82 tons/yr). Mercury emissions occur when mercury in the sewage is combusted at high
temperatures, vaporizes and exits through the gas exhaust stack. Landfilled sludge or sludge applied to
farmland are also potential sources of mercury emissions. These sources are not addressed in this
inventory.
A total of 206 SSIs currently operate in the United States. An estimated 1.5 x 106 Mg
(1.65 x 106 tons) of sewage sludge on a dry basis are incinerated annually (U.S. EPA, 1993b).
Table A-13, Appendix A shows the geographic distribution of sewage sludge incinerators throughout
the country. 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 32, 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: clewatering 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.
The 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 10'5 Ib/ton) for multiple hearth combustors controlled with a combination
of venturi and impingement scrubbers, 0.03 g/Mg (6.0 x 10"5 Ib/ton) for fluidized bed combustors
controlled with a combination of venturi and impingement scrubbers, 2.3 g/Mg (4.6 x 10"3 Ib/ton) for
multiple hearth combustors controlled with a cyclone scrubber, 1.6 g/Mg (3.2 x 10"3 Ib/ton) for
multiple hearth combustors controlled with a combination of cyclone and venturi scrubbers, and
0.97 g/Mg (1.94 x 10"3 Ib/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"5 Ib/ton) for SSIs controlled with a combination of venturi and impingement scrubbers and
1.623 g/Mg (3.25 x 10"3 Ib/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, 1.5 x 106 Mg (1.65 x 106 tons), by the appropriate percentage and
emission factor gives a mercury emission estimate of 0.009 Mg/yr (0.01 tons/yr) for SSIs controlled
with a combination of venturi and impingement scrubbers and an estimate of 1.64 Mg/yr (1.81 tons/yr)
for SSIs controlled by some other means. The overall mercury emissions estimate from SSIs is, thus,
1.65 Mg/yr (1.82 tons/yr).
June 1996 4-29 SAB REVIEW DRAFT
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4.1.7 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 1991, there were
400,465 cremations in the slightly more than 1,000 crematories located throughout the United States
(Cremation Association of North America, 1992). Table A-8 in Appendix A lists the estimated
number of crematories located in each State and the estimated number of cremations performed.
Information was not available on the location of individual crematories (Vander Most and Veldt,
1992).
No data are available for the average quantity of mercury emitted for a cremation in the
United States. Three estimated levels have been cited for European countries (Switzerland, Germany
and the United Kingdom) with an estimated emission rate of 1 gram of mercury per cremation
recommended as a typical value (Springer, 1993). This emission factor may not be applicable to
cremations in the United States because dental care programs in the United States differ markedly
from those in Europe. Consequently, the average number of mercury amalgam fillings per person may
differ considerably, with Europeans believed to have more fillings per person than Americans.
Because the average number of fillings per person and the average mercury content per filling have a
direct impact on the estimated mercury emissions, this European emission factor may not provide an
accurate estimate of mercury emissions from cremations in the United States. Multiplying the
European emission factor of 1 gram of mercury per cremation by the total number of U.S. cremations
in 1991 (400,465) gives a mercury emission estimate of 0.4 Mg/yr (0.4 tons/yr).
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 1011 Btu/hr) (U.S. EPA, 1982). No specific data on the
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).
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
June 1996 4-30 SAB REVIEW DRAFT
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combustors, but results for other combustion sources suggest that efficiencies will be low, probably
50 percent or less (U.S. EPA, 1993a).
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 0.34 x 10'5 kg/Mg (0.67 x 10"5 Ib/ton) of
wood burned 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 1011 Btu/hr),
by the average 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, 1982). 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 0.3 Mg/yr (0.33 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.
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"2 Ib/ton) is 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 Hazardous Waste Combustors
For the purpose of this emissions inventory, hazardous waste combustors include hazardous
waste incinerators and lightweight aggregate kilns. Although hazardous waste burning cement kilns
are also typically classified as hazardous waste combustors, mercury emissions from cement
manufacturing are considered separately (see Section 4.2.4).
Based on the U.S. EPA's 1995 emission estimates (U.S. EPA, 1995b), hazardous waste
incinerators and lightweight aggregate kilns currently combine to emit a total of 4.5 Mg/year (5.0
tons/year) of mercury. Of this amount, hazardous waste incinerators are estimated to emit 4.3 Mg/year
(4.7 tons/year), or approximately 95 percent of the total, while lightweight aggregate kilns are
estimated to emit 0.25 Mg/year (0.27 tons/year), or about 5 percent of the total.
June 1996 4-31 SAB REVIEW DRAFT
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4.1.9.1 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 faculties 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., HC1, Cy
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 been used on any full-scale incinerators hi the U.S.
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 afterburner, operating at 100-200 percent excess air and 2000-2500°F, where unbuipt components of
the kiln flue gas 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 840-1SOO°F. Further reaction occurs in the volume above
the bed at temperatures up to 1800°F.
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 unbumt
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.
June 1996 4-32 - SAB REVIEW DRAFT
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4.1.9.2 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.
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 SO2, CO2, SO3, and Oj) that eventually trigger the desired bloating action within the material. As
temperatures reach their maximum (approximately 2100°F), the semiplastic raw material becomes viscous and
entraps the expanding gases. 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.
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.
4.2 Manufacturing Sources
Manufacturing sources, including processes that use mercury directly and those that produce
mercury as a byproduct, account for an estimated 29.1 Mg/yr (32 tons/yr) of mercury emissions
generated in the United States. Emissions from these sources are presented in Table 4-10 and are
discussed below.
4.2.1 Primary Lead Smelting
Primary lead smelters recover lead from a sulfide ore, which may contain mercury. The
smelters emitted an estimated 8.2 Mg (9 tons) of mercury into the atmosphere in 1990. Table 4-11
lists the locations and 1990 production rates of the three primary lead smelters that are currently
operating in the United States; the locations of these smelters are displayed in Figure 4-9.
June 1996
4-33 •
SAB REVIEW DRAFT
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Table 4-11
1990 U.S. Primary Lead Smelters and Refineries*
Smelter
ASARCO, East Helena, MT
ASARCO, Glover, MO
Doe Run (formerly St. Joe)
Refinery
ASARCO, Omaha, NE
ASARCO, Glover
Doe Run, Herculaneum, MO
1990 Lead Production
Tons (Megagrams)
65,800 (72,500)
112,000 (123,200)
231,000 (254,100)
a Woodbury, 1992.
Figure 4-9
Primary Lead Smelters
June 1996
4-35
SAB REVIEW DRAFT
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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.
Primary lead smelters use high-efficiency emission control systems to reduce the levels of PM
and SO2 from the blast furnace and sintering machines. Centrifugal collectors (cyclones) are
used in conjunction with baghouses or ESPs for PM control. Control of SO2 emissions from sintering
is 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 paniculate 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 dressing area because these processes take place at high temperatures.
No recent mercury emission factors are available for the three currently operating primary lead
smelters. 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). These emission
factors were used in the preliminary analysis to evaluate the potential mercury emissions from primary
lead smelting. Based on these emission factors and primary lead production estimates, EPA estimated
total mercury emissions at 8.2 Mg (9 tons) for primary lead smelters. This estimate is reflected in this
document as the 1990 estimate and was used as the basis for subsequent analyses in this Report (e.g.,
the long-range transport analysis). Because numerous assumptions were used to convert the emission
factors to a lead production basis, this emission estimate was considered to have a high degree of
uncertainty.
Due to the relatively high emission potential identified for primary lead smelters, EPA has
recently reviewed the emission estimates. The review determined that the 1990 estimate overstated
June 1996 4-36 SAB REVIEW DRAFT
-------
potential emissions. As stated above, the 1990 estimate was based on data provided for the El Paso
smelter. This smelter obtained lead ore from several sources from both within and outside of the
United States. These ores had a variable mercury content depending upon the source of the ore. Two
of the three current smelters are not custom smelters; they typically process concentrates from
southeast Missouri. These concentrates have a very low mercury content. The ASARCO-East Helena
plant, although a custom smelter, also processes low mercury concentrates. In addition, the earlier El
Paso emissions data included emissions from processes that do not take place at the three operating
primary lead smelters. Based on the findings of this review, the U.S. EPA's current estimate of
national mercury emissions from primary lead smelters is about 1.3 Mg/yr (1.5 tons/yr).
4.2.2 Secondary Mercury Production
Secondary mercury production (mercury recycling) involves processing scrapped mercury-
containing products, industrial waste and scrap, and scrap mercury from government stocks.
Secondary mercury production is estimated to have accounted for approximately 6.7 Mg (7.4 tons) of
mercury emissions in 1991. Major sources of recycled mercury include dental amalgams, scrap
mercury from instrument and electrical manufacturers (lamps and switches), phosphor from discarded
fluorescent lamps, wastes and sludges from research laboratories and electrolytic refining plants, and
mercury batteries (U.S. EPA, 1993a). Table 4-12 lists the five major companies that were involved in
secondary mercury production in 1989.
Table 4-12
1989 U.S. Mercury Recyclers*
Adrow Chemical Company Wanaque, NJ
Bethlehem Apparatus Company, Inc. Hellertown, PA
D. F. Goldsmith Chemical and Metals Corp. Evanston, IL
Mercury Refining Company, Inc. Latham, NY
Wood Ridge Chemical Company Newark, NJ
a Bureau of Mines, 1991.
During secondary mercury production, emissions may potentially occur from the retort or
furnace operations, the distillation process and the discharge to the atmosphere process (Reisdorf and
D'Orlando, 1984; U.S. EPA, 1984). The major mercury emission sources are the condenser exhaust
and fugitive vapor emissions that occur during unloading of the retort chamber. Mercury emissions
can also occur in the filling area if the flask overflows and during the bottling process.
Information on specific emission control measures is limited and site-specific. If a scrubber is
used, mercury vapor or droplets in the exhaust gas may be removed by condensation in the spray. No
information was found for other control measures that are being used in secondary mercury production
processes. Concentrations 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
methods are used in conjunction with any type of emission control device (Reisdorf and D'Orlando,
1984). 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.
June 1996 4-37 SAB REVIEW DRAFT
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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, 1993a). 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 538°C (1000°F) 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 [HO], SO2). From the aqueous
scrubber, the 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 Ib) or 2.3 kg (5 Ib) 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).
Because the secondary mercury production process has not been recently tested, virtually no
data are available for this process. In 1973, emission factors were estimated to be 20 kg of mercury
emitted per megagram of mercury processed (40 Ib/ton) due to uncontrolled emissions over the entire
process (Anderson, 1973). These data have a high degree of uncertainty associated with them for the
following reasons: (1) they are 20 years old, and processes may have changed substantially since they
were generated; and (2) no information is available on their bases.
In 1991, 122 Mg (134 tons) of mercury were recycled from industrial scrap and 215 Mg
(237 tons) from government stocks in the United States (Josinski, 1992). This total (337 Mg
[371 tons]) does not include in-house mercury reclamation at industrial plants using mercury.
Multiplying the total mercury recycled by the emission factor gives a mercury emission estimate of
6.7 Mg (7.4 tons) for 1991. Again, this estimate has a high degree of uncertainty because of the
limited reliability of the emission factor.
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).
June 1996 4-38 SAB REVIEW DRAFT
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Mercury is recovered from crushed lamps by heating the crushed material to vaporize the
mercury and then cooling the offgas 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).
As presented previously in Figure 3-1,2 percent of fluorescent lamps are estimated to be
recycled each year. 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 et al., 1993). This number may be
increasing due to the increasing number of recycling plants. Most lamp recycling facilities are located
in California and Minnesota (Battye et al., 1994).
4.2.3 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 hi 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. The chlor-
alkali industry, however, is gradually 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 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 4-13 lists U.S. mercury-cell chlor-alkali production facilities and then1 capacities. Figure
4-10 shows the location of these facilities across the U.S. The chloj-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).
With the downward trend of chlor-alkali production, 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
June 1996 4-39- SAB REVIEW DRAFT
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Table 4-13
1991 U.S. Mercury-Cell Chlor-Alkali Production Facilities9
Facility
Georgia-Pacific Corp., Chemical
Division
BF Goodrich, Chemical Group
Hanlin Group, Inc., LCP
Chemicals Division
ASHTA Chemicals, Inc.
Occidental Petroleum
Corporation, Electrochemicals
Division
Olin Corporation, Olin
Chemicals
Pioneer Chlor Alkali Company,
Inc.
PPG Industries, Inc., Chemicals
Group
Vulcan Materials Company,
Vulcan Chemicals Division
Location
Bellingham, WA
Calvert City, KY
Reigelwood, NC
Orrington, ME
Ashtabula, OH
Deer Park, TX
Delaware City, DE
Muscle Shoals, AL
Augusta, GA
Charleston, TN
St Gabriel, LA
Lake Charles, LA
New Martinsville, WV
Port Edwards, WI
TOTAL
Capacity,
103 Mg/yr
82
109
48
76
' 36
347
126
132
102
230
160
233
70
65
1,816
Capacity,
103 tons/yr
90
120
53
80
40
383
139
146
112
254
176
256
77
72
1,998
1991
emissions11
Ib/yr
200
1,206
528
735
N/A
1,290
532
184
1,540
1,892
1,240
1,440
1,085
1,030
12,902
(5,858 kg/yr)
N/A = Not available from survey questionnaires. It is assumed that facilities not reporting mercury emissions
produce no mercury emissions.
* SRI International 1991
b BF Goodrich, 1992; Georgia-Pacific, 1993; LCP Chemicals, 1993a; LCP Chemicals, 1993b; Occidental, 1993a;
Occidental, 1993b; Occidental, 1993c; Olin Chemicals, 1993a; Olin Chemicals, 1993b; Pioneer Chlor-Alkali, 1993; PPG
Industries, 1993a; PPG Industries, 1993b; Vulcan Materials, 1993
June 1996
4-40
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Figure 4-10
Chlor-Alkali Production Facilities
The Mercury-Cell Chlor-Alkali Process
The mercury-cell chlor-alkali process consists of two electrochemical cells, 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).
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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).
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 127°C (200 to 260°F) 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 43°C (90 to 110°F). 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 13°C (37 to
55°F). 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 hi 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. The reported mercury emissions were 5.9 Mg (6.5 tons) and included 13 of the
14 mercury cell chlor-alkali production facilities listed in Table 4-=ll. Data are also available from the
Toxic Release Inventory (TRI) (U.S. EPA, 1992e). Those facilities not reporting mercury emissions in
June 1996 4-42 SAB REVIEW DRAFT
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the section 114 questionnaire responses or in the 1991 TRI were assumed to produce no mercury
emissions because it is a Federal law to report such emissions.
4.2.4 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 106 Mg (81 x 106 tons) (U.S. EPA, 1993a). Of this total, 201
kilns were active and had a total clinker capacity of 71.8 x 106 Mg (79.1 x 106 tons) (U.S. EPA,
1993a). 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 5.9 Mg (6.5 tons). 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 Rgure 4-11.
€
Figure 4-11
Cement Manufacturing Plants
The primary sources of mercury emissions from portiand 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 nulling 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.
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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., kirn). 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 hi 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 hi 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 hi coal and
oil and may also be present hi appreciable quantities in the waste-derived fuels mentioned above. Because
mercury evaporates at approximately 350°C (660°F), 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 finished portland cement
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.
Mercury present in the raw material and the fuel is likely to be emitted from all four cement
processes summarized in the box on the next page. 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
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these were from the same kiln). This information included clinker production as well as mercury
emission rates and process conditions.
Although the emission tests were COC tests, the data used to calculate the emission factor
were taken from the "background" test runs when the usual fuel was being fired at normal operating
conditions. If abnormal operating conditions were noted in the test report (e.g., hotter than normal
clinkering), then the test results were not used. In some test reports, the mercury emissions were
similar when either coal or WDF was being fired If the emissions from the WDF burn was no higher
than the emissions when coal alone was fired, the emissions data across test runs were averaged.
The calculated emission factors ranged from 0.0023 to 0.49 g/Mg of clinker (4.5 x 10"6 to 9.7
x 10"4 Ib/ton). The average emission factor for the 15 test runs was 0.087 g/Mg of clinker (1.7 x 10"4
lb/ton)(U.S. EPA, 1993a).
Kilns firing WDF or coal only may have different mercury emissions than the average
emission factor would indicate. As a result, the emissions estimated here for the portland cement
industry as a whole (i.e., including kilns that fire fossil fuel or WDF) may underestimate or
overestimate emissions as the mercury content of both WDF and coal may vary. Seven percent of the
entire fuel consumption of the cement manufacturing industry in 1990 consisted of WDF. Coal was
the largest fuel source, at over 71 percent of consumption, followed by coke at over 13 percent of
consumption (Carrol, 1994).
The total production of portland cement in 1990 was 67 J x 106 Mg (74.5 x 106 tons)
(95.7 percent of the total cement production) (U.S. EPA, 1993a). Of the total production of portland
cement, 96 percent was clinker, and the remaining 4 percent was other ingredients (U.S. EPA, 1993a).
Multiplying the total amount of clinker produced in 1990 (total production of portland cement
multiplied by 96 percent) by the above average emission factor gives an estimate of 5.9 Mg (6.5 tons)
of mercury emissions from portland cement manufacturing for 1990. This emissions estimate is
expected to be updated in 1995 during development of U.S. EPA's Combustion Strategy as well as for
rulemaking for this source category under section 112 of the CAA.
Industry estimates for this category (based on a review of the same test data) are 2.9 Mg (3.3
tons) of mercury emitted in 1990. This estimate is calculated using an estimated average emission
factor of 9.1xlO"5 Ib of mercury per ton of clinker (Portland Cement Association, 1995).
4.2.5 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 4-14
gives the locations and 1992 production capacities of primary copper smelters currently operating in
the United States; these smelter locations are displayed in Figure 4-12.
Copper smelters use high efficiency air pollution control options to control PM and SO2
emissions from smelting furnaces and converters. Electrostatic precipitators are the most common PM
control device at copper smelters. Control of SO2 emissions is achieved by absorption to sulfuric acid
in the sulfuric acid plants, which are common to all copper smelters.
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Table 4-14
1992 U. S. Primary Copper Smelters and Refineries
Smelter
ASARCO Inc.
Cyprus Miami Mining Co.
BHP Copper Co.
Copper Range Co.a
Phelps Dodge \
Chino Mines Co.
ASARCO Inc.
Kennecott
Location
Hayden, AZ
Globe, AZ
San Manuel, AZ
White Pine, MI
Hidalgo, NM
Hurley, NfcT
H Paso, TX
Garfield, UT
1992 Capacity, Mg (tons)
190,900 (210,200)
180,000 (198,000)
290,000 (319,000)
60,000 (66,000)
190,000 (209,000)
170,000 (187,000)
104,000(114,400)
210,000 (231,000)
Ceased operations in February 1995.
Figure 4-12
Primary Copper Smelters
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The Primary Copper Smelting Process
The copper smelting process sequentially involves drying ore concentrates, smelling 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 SO2 to be processed in the
sulfuric acid plant
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 2500°C (4530°F). AAy
mercury containedVn the concentrate will likely be emitted during the flash smelter process step and directed to the
sulfuric acid plant (Buonicore and Qavis, 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
T imitftH 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 Cv^S, called '"white metal" accumulates
in the bottom of the converter. A renewed air blast then oxidizes the copper sulfide to SO2, leaving blister copper.
Blister copper is then removed and transferred to refining facilities. Further purification may involve fire refining and
electrolytic refining (Buonicore and Davis, 1992).
Mercury emission data for primary copper smelting facilities are very limited. One emission
test report for the Copper Range Company located in White Pine, Michigan containing metals analysis
results was available (TRC Environmental, 1992). Before ceasing operations in February 1995, this
facility operated a reverberatory furnace with an ESP to control PM. The exhaust stream from the
converter (which is uncontrolled) was mixed with the exhaust from the ESP outlet and was routed
through the main stack and discharged into the atmosphere. Testing for metals was performed at a
location in the main stack downstream from the point where the two exhaust streams (from the ESP
outlet and the convenor) are combined. Mercury emissions were measured for three modes of
convenor operation: slag-blow, copper-blow and converter idle (no blow) cycles. The mercury level
during the convenor idle cycle was measured to be the highest, corresponding to a mercury emission
rate of 0.0753 kg/hr (0.166 Ib/hr). During the slag-flow and copper-flow periods, the emission rates
were 0.0494 kg/hr (0.109 Ib/hr) and 0.0635 kg/hr (0.140 Ib/hr), respectively. Additionally, the plant
capacity was reported to be approximately 38 Mg/hr (42 tons/hr) of feed, which consists of mill
concentrate, limestone, iron pyrites and recycled material. The actual process rate during the test is
not known, so an emission factor cannot be calculated from this test.
Because the feed mix varies from facility to facility and because the Copper Range Company
is the only facility in the United States that operated a reverberatory furnace without any acid plants,
the emission data from the Copper Range Company are not representative of current industry practice.
A mercury emissions estimate of 0.6 Mg/yr (0.7 tons/yr) from this one facility was, therefore,
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calculated using a weighted average of the emission rates for the three modes of converter operation
(0.068 kg/hr [0.15 lb/hr]), and an operating schedule of 8,760 hr/yr. The Copper Range facility also is
no longer active.
A more recent analysis of the seven copper smelters currently operating in the U.S. has been
performed since the 1990 emissions estimate described above. Mercury emission rates from these
seven smelters are presented in Table 4-15 along with the mercury concentration of ore. These data
show that emissions range from less than 1 Ib/year to 40 Ibs/year. These emission rates are based on
both stack testing and engineering judgment. As a result, the U.S. EPA estimates 1995 nationwide
mercury emissions from primary copper smelters to be about 0.09 Mg/year (0.08 tons/year).
Table 4-15
Mercury Ore Concentrate and Emissions from
Primary Copper Smelters in the U.S.
Smelter
ASARCO - El Paso
ASARCO - Hayden
Copper Range
Cyprus Miami
Kennecott
BHP Copper Co.
Phelps Dodge-Hidalgo
Phelps Dodge-Chino
Mercury in Ore
Concentrate
Ib/yr
1,769
2,444
940
CBIa
NAa
2,240
5,768
585
Mercury
Emissions
Ib/yr
1.8
35
1,951
34
35
•
40
0.09
7.5
Basis of
Emission Values
Emissions Test
Emissions Test and
Engineering Judgment
Emissions Test
Emissions Test
Emissions Test and
Engineering Judgment
Emissions Test and
Engineering Judgment
Engineering Judgment
Engineering Judgment
' CBI means Confidential Business Information that is unavailable to the public. NA means not available.
4.2.6 Lime Manufacturing
Lime is produced in various forms, with the bulk of production yielding either hydrated lime
or quicklime. In 1992, producers sold 16.4 x 106 Mg (18 x 106 tons) of lime produced at 113 plants
in 32 States and Puerto Rico. The 1992 production represented a 4 percent increase over 1991
production. The leading domestic uses for lime include steelmaking, flue gas desulfurization, pulp and
paper manufacturing, water purification, and soil stabilization (Miller, 1993).
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Table 4-16 identifies the top 10 lime-producing plants in the United States, in order of total
output for 1991 (Hammond, 1993). 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).
Table 4-16
Leading 1991 UJS. Lime Producing Plants3
Plant
Ste. Genevieve
Maysville Division
Black River Division
Montavello Plant
Woodville Plant
Longview Division
South Chicago Plant
Nelson Plant
Clifton Plant
Annville Plant
Company
Mississippi Lime Company
Dravo Lime Company
Dravo Lime Company
Allied Lime Company
Martin Marietta Magnesia
Dravo Lime Company
Marblebead Lime Company
Chemstar, Inc.
Chemical Lime, Inc.
Wimpey Minerals PA, Inc.
Location
Ste. Genevieve County, MO
Mason County, KY
Pendleton County, KY
Shelby County, AL
Sandusky County, OH
Shelby County, AL
Cook County, IL
Yavapai County, AZ
Bosque County, TX
Lebanon, PA
11 Hammond, 1993.
Commercial production of the various forms of lime involves the following basic steps.
•
1. Quarrying raw limestone (or limestone and dolomite);
2. Stone processing or crushing and sizing in preparation for calcination;
3. Calcining the crushed stone in high temperature kilns (producing quicklime);
4. Hydrating the processed lime (to produce hydrated lime from quicklime); and
5. Miscellaneous transfer, storage and handling processes.
Emissions from quarrying or stone extraction are largely restricted to fugitive dust, as are the
emissions from stone processing, crushing and sizing. Mercury emissions are expected to be
negligible from these initial steps in lime production.
Calcining, which uses high-temperature kirns to convert carbonate to oxide (removing C02), is
the lime production step from which most mercury emissions are expected. Rotary kilns are primarily
used in the calcining step in the United States, accounting for 90 percent of domestic lime kilns.
Other types of lime kilns include vertical kilns, rotary hearths and fluidized bed kilns. During
calcination, kiln temperature may reach as high as 1820°C (3300°F) (U.S. EPA, 1993a). Because
mercury has a boiling point of 350°C (660°F), most of the mercury that exists as impurities in the
processed stone will likely emit as an air pollutant during calcination.
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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 rne 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)2). The hydration step may be immediately preceded by some crushing,
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.
Representative estimates of mercury emissions from lime manufacturing are not possible based
on the available data from lime kilns around the country. An ongoing U.S. EPA study to update
AP-42, Section 8.15, on lime manufacturing emission factors has reviewed and summarized test data
for lime calcining at 93 kilns (U.S. EPA, 19921). Pollutants identified and noted in a summary of the
test data did not indicate any mercury emissions and gave little or no indication that emissions tests at
lime kilns have sampled and analyzed for trace metals. Two very limited estimation efforts for
mercury emissions are offered in the following discussion: one using 1983 mercury emission test data
from five Wisconsin lime plants; and, the other from a draft report on 1983 mercury emission data
from a pulp and paper lime kiln, which is referenced in SPECIATE.
An emission estimate based on mass balances generated from information for mercury content
in limestone from the five operating lime kilns in Wisconsin in 1983 was 18 kg/yr (39 Ib/yr) for all
the kilns combined (Bureau of Air Management, 1986). In 1983, these five lime plants produced
0.29 x 106 Mg (0.32 x 106 tons) of lime (Miller, 1993b). Assuming uniform emissions for each ton of
production suggests that 5.53 x 10"5 kg (1.22 x 10"4 Ib) of mercury were emitted for each Mg (ton) of
lime produced. These data do not account for any differences in fuel used to heat the kilns or any
differences in raw materials used. Natural gas, which is believed to contain negligible amounts of
mercury, is used to fire 33 percent of the lime kilns. Total estimated annual emissions would,
therefore, be reduced by 33 percent to reflect the lack of mercury emissions from natural gas.
If the Wisconsin data are extrapolated to the lime production in the United States in 1992, an
annual estimate of mercury emissions from lime kilns of 0.91 Mg/yr (1.00 ton/yr). Assuming that
33 percent of lime kirns use natural gas as their fuel source and produce no mercury emissions reduces
this estimate to 0.6 Mg/yr (0.7 tons/yr). This estimate has a high level of uncertainty because material
composition could vary significantly across the country, and the fuel type(s) used in Wisconsin may
not be representative of those used nationally.
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4.2.7 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 0.5 Mg (0.5 ton) in 1992. No information on locations
of manufacturers of electrical apparatus that specifically contain mercury is available.
4.2.7.1 Electric Switches
Hie 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.
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 die 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, custom-designed to fit specific equipment, can reduce mercury vapor and mercury PM
(Reisdorf and D'Orlando, 1984).
The amount of mercury used for the manufacture of switches and thermostats decreased 50
percent from 155 tons in 1989 to 77 tons in 1990. 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
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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 4-17 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 should 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.
Table 4-17
Discards of Mercury in Electric Switches9
Year
1987
1988
1989
1995
2000
Electric Switch Production
1,000,000
1,000,000
1,000,000
1,000,000
1,000,000
Weight of Mercury in
Switches (tons)
3.9
3.9
3.9
3.9
3.9
Weight of Mercury Discarded in
MSW (tons)
0.39
0.39
0.39
1.93
1.93
a U.S. EPA, 1992a.
4.2.7.2 Thermal Sensing Instruments and Tungster 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
June 1996 4-52 SAB REVIEW DRAFT
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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
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, 1993a).
4.2.7.3 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 me 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).
4.2.7.4 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 550 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 a small amount (15 to 250 mg [3.3 x 10"5 to 5.5 x 10"4 lb]) of mercury is
added. Some of the mercury combines with the emulsion on the interior of the bulb over its life. 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.
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The names and division headquarters of the fluorescent lamp manufacturers in the United
States in 1995 are shown to Table 4-18 (U.S. EPA, 1993a).
Table 4-18
1992 U.S. Fluorescent Lamp Manufacturers' Headquarters9
Company
Duro-Test Corp.
General Electric
OSRAM Corp.b
Philips Lighting Company
Division headquarters
North Bergen, NJ
Cleveland, OH
Montgomery, NY
Somerset, NJ
a U.S. EPA, 1993a.
b National Electrical Manufacturers Association, 199S.
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).
4.2.7.5 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 hi 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 Ib/ton) hi 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.
In 1992, 104 Mg (116 tons) of mercury were used hi all electrical apparatus production
(36 Mg [40 tons] for electric lighting and 69 Mg [76 tons] for wiring devices and switches) (Bureau of
Mines, 1992).3 Multiplying the emission factor above by the 1992 usage gives a mercury emission
estimate of 0.42 Mg (0.46 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.
3 The Bureau of Mines previously reported mercury use for electric lighting to be 61 tons. This number is
being revised to an estimated 40 tons (National Electric Manufacturers Association, 1995).
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4.2.8 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, mis 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).
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 200°C (390°F), 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 (buming-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).
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, 1993a). 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, 1993a).
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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 landfilled 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 Ib/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).
4.2.9 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 1991, mercury
emissions from carbon black production were estimated to be 0.23 Mg (0.25 ton). This estimate is
expected to be an overestimate because it is based on production capacity and not on actual
production. Table 4-19 lists the names, locations and annual capacities of U.S. producers of carbon
black in 1991 (SRI International, 1992).
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/Nm3) (0.003 gr/dscf). Mercury emissions
from 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 170°C (325°F), the fabric filter may capture a significant fraction of the condensed
mercury, thus providing some degree of emission control (Taylor, 1992).
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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 ISO 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 burn 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).
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"4 Ib/ton) from the main process vent (Serth and
Hughes, 1980). 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 1991, the total capacity for carbon black production was 1.55 x 1()6 Mg (1.7 x 106 tons)
(SRI International, 1992). Multiplying the total capacity by the emission factor above gives a mercury
emission estimate of 0.23 Mg (0.25 tons). This estimate may be greater than the actual emissions
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.
4.2.10 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 mercury-
zinc) 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 (HgCl2), or
mercurous chloride (Hg2Cl2) (White and Jackson, 1993). As a result of technological improvements
made by the 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 4-20 lists the
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Table 4-19
1992 U.S. Carbon Black Production Facilities3
Company
Cabot Corporation North
American Rubber Black
Division
Chevron Corporation,
Chevron Chemical Company,
subsidiary, Olevins and
Derivatives Division
Degussa Corporation
Ebonex Corporation
General Carbon Company
Hoover Color Corporation
J.M. Huber Corporation
Phelps Dodge Corporation
Colombian Chemical Company,
subsidiary
Sir Richardson Carbon &
Gasoline Company
Witco Corporation
Continental Carbon Company,
subsidiary
Location
Franklin, Louisiana
Pampa, Texas
Villa Platte, Louisiana
Waverly, West Virginia
Cedar Bayou, Texas
Arkansas Pass, Texas
Belpre, Ohio
New Iberia, Louisiana
Melvindale, Michigan
Los Angeles, California
Hiwassee, Virginia
Baytown, Texas
Borger, Texas
Orange, Texas
El Dorado, Arkansas
Moundsville, West Virginia
North Bend, Louisiana
Ulysses, Kansas
Addis, Louisiana
Big Spring, Texas
Borger, Texas
Phoenix City, Alabama
Ponca City, Oklahoma
Sunray, Texas
Type of
processb
F
F
F
F
A
F
F
F
C
C
C
F
FandT
F
F
F
F
F
F
F
F
F
F
F
TOTAL
Annual capacity0
K^Mg
141
32
127
82
9
59
64
109
4
0.5
0.5
102
79
61
50
77
109
36
66
52
98
27
114
45
1,546
103 tons
155
35
140
90
10
65
70
120
4
0.5
0.5
112
87.5
67.5
55
85
120
40
72.5
57.5
107
30
125
50
1,700
a SRI International, 1992.
A = acetylene decomposition; F = furnace; C = combustion; T = thermal.
c Capacities are variable and based on SRI estimates as of January 1, 1992.
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Table 4-20
1992 U.S. Mercuric Oxide, Alkaline Manganese, or
Zinc-Carbon Button Cell Battery Manufacturer!;9
Manufacturer
Alexander Manufacturing Company
(AMC, Inc.)
Duracell, USA
Eagle-Picher Industries, Inc.
Eveready Battery Company, Inc.
Mutecd
Rayovac Corp.
Production site
Mason City, IA
Cleveland, TN
LaGrange, GA
Lancaster, SC
Lexington, NC
Colorado Springs, CO
Maryville, MO
Red Oak, IA
Fremont, OH
Bennington, VT
Asneboro, NC (2 plants)
Columbus, GA (Corporate offices)
Madison, WI
Fennimore, WI
Portage, WI
1990 Mercury TRI emissions
kg (lb)b
0(0)
NRC
NR
9(20)
3(70)
NR
14 (30)
NR
NR
1(2)
2(5)
NR
0(0)
5(10)
NR
* U.S. EPA, 1993a.
b U.S. EPA, 1992e.
0 NR = Not reported, company did not report mercury emissions in 1990 TRI.
Mutec is a joint venture between Eastman Kodak and Panasonic.
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 hi 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).
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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, palletizing, 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
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 Ib) 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 4-21 (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 Ib) in the 1990 TRI (U.S. EPA, 1992e).
In 1992, 16 Mg (18 tons) of mercury were used in the production of batteries in the United
States (Bureau of Mines, 1992). Multiplying the mercury usage by the emission factor developed for
the facility in Wisconsin gives a mercury emission estimate of 0.02 Mg (0.02 tons) for 1992. This
estimate is highly uncertain, however, because of the concerns discussed above about the reliability of
the emission factors (U.S. EPA, 1993a). 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.
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Table 4-21
Emission Source Parameters for an Integrated
Mercury Button Cell Manufacturing Facility3
Building/source description13
Emission rate0
g/d
Ib/d
Exit temp. (K); control device
Main plant
Control room
1. Blending, slugging,
compacting, granulating
2. Slugging, granulating
3. Pelleting, consolidating
4. Pelleting, consolidating
4a. Pelleting, consolidating
5. Blending, compacting,
granulating, pelleting,
consolidating
Anode room
6. Amalgam, dewatering
6a. Vacuum dryer
6b. Blending
7. Pelleting, zinc amalgam
6.12
1.22
1.63d
42.46
6.53
1.36d
0.0135
0.0027
o.oose*1
0.0936
0.0144
0.003b
297; Baghouse
297; Baghouse
295; Baghouse
297; Baghouse
297; Baghouse
297; Baghouse
1.82d
0.46d
0.91d
4.08d
Cell assembly area
8. Assembling calls
28.58
0.004d
0.001d
0.002d
0.009d
0.0630
297; Uncontrolled
297; Uncontrolled
297; Uncontrolled
295; Baghouse
295; Baghouse for PM. Vapor by
recirculating air through prefilters
and charcoal filters
a U.S. EPA, 1984.
b Source names are those used by facility.
0 Emission rates were measured by facility except where noted.
d Estimated emission rate by facility.
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4.2.11 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, 1993a).
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 1991, nine U.S. gold mines (seven 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 4-22. 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.
Table 4-22
1992 U.S. Byproduct Mercury-Producing Gold Mines3
Mine
Alligator Ridge
Getchell
Carlin Mines Complex
Hog Ranch
Jetritt Canyon (Enfield Bell)
McLaughlin
Mercur
Paradise Peak
Pinson and Kramer Hill
County/State
White Pines, NV
Humboldt, NV
Eureka, NV
Washoe, NV
Elko, NV
Napa, CA
Tooele, UT
Nye, NV
Humboldt, NV
Operator
USMX, Inc.
FirstMiss Gold
Newmont Gold Co.
Western Hog Ranch Co.
Independence Mining Co., Inc.
Homestake Mining Co.
Barrick Resources (USA) Inc.
Permanently closed in September 1993
Pinson Mining Co.
a Bureau of Mines, 1994.
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, 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.
June 1996
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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 en 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 exhausl. gas then passes through a
lime sulfur dioxide (SO^) scrubber prior to 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 ;i 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, SO2, PM, water vapor, and other volatile components,
passes 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 SO2
scrubber prior to 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 SO2. If the mercury concentration
in the 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 SO2; if the SO2
concentration is low, a 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 SO2 prior to discharging to the
atmosphere.
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).
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.
No specific data on emission factors from potential sources of mercury emissions from
mercury ore mining have been published since 1973 (U.S. EPA, 1993a). The 1973 report gives a total
emission factor of 0.171 kg of mercury emitted for each megagram of mercury ore mined (0.342
June 1996
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Ib/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.
4.2.12 Mercury Compounds Production
The production of mercury compounds presents a potential source of mercury emissions into
the atmosphere. Common mercury, compounds include mercuric chloride, mercuric oxide, and
phenylmercuric acetate (PMA). Table 4-23 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, 1993a).
During the production of mercury compounds, emissions of mercury vapor and paniculate
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, 1993a).
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 hi Table 4-23, the only company that reported significant emissions (227 kg [500 lb]) in the
1991 TRI was Mallinkrodt, Inc.
4.2.13 Byproduct 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 4-24 lists U.S. byproduct coke oven facilities in 1991 (Huskanen, 1991) and Figure
4-13 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).
June 1996 ' 4-64 SAB REVIEW DRAFT
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Table 4-23
1991 U.S. Mercury Compound Producers3
Producer
Atochem North America, Inc.,
Chemical Specialties Division
Atomergic Chemetals Corp.
Cambrex Corp., CasChem, Inc.,
Subsidiary (formerly Cosan
Chem. Corp.)
W.A. Cleary Corp.
Deepwater, Inc.
GFS Chemicals, Inc.
Hills America, Inc.
Imsera Group, Inc.,
Mallinkrodt Inc.,
Subsidiary, Mallinkrodt Specialty
Chem. Co.,
Morton International, Inc.,
Specialty Chemicals Group,
Advanced Materials, CVD Inc.
Subsidiary
Polychemical Laboratories, Inc.
R.S.A Corporation
Troy Chemical Corp.
Location
Tulsa,OK
Fanningdale, NY
Caristadt, NJ
Somerset, NJ
Carson, CA
Columbus, OH
Elizabeth, NJ
Erie, PA
Wobum, MA
Melville, NY
Ardsley, NY
Newark, NJ
1991 TRI
emissions,
kg (lb)b .
NRC
NR
18(40)
NR
NR
NR
0(0)
227 (500)
NR
NR
NR
0(0)
Compounds)
HgF2
Thimerosal (Merthiolate)
Phenylmercuric acetate
(PMA), Phenylmercuric
oleate:
Phenylmercuric acetate
(PMA)
HgI2
HgBr2, HgI2, Hg(N03)2,
HgS04
Phenylmercuric acetate
(PMA)
HgCl2 on carbon support
(catalyst for vinyl chloride
manufacture)
Highly purified
dimechylmercury, (CH3)2Hg,
for diemical vapor deposition
(CVD) of thin films
Thimerosal (Merthiolate)
Hg(SCN)2
Phenylmercuric acetate
(PMA)
a SRI International, 1991.
b U.S. EPA, 1992e.
0 NR = Not reported; company did not appear in 1991 TRI.
June 1996
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Table 4-24
1991 U.S. Byproduct Coke Producers*
Facility
Acme Steel, Chicago, IL
Armco, Inc., Ashland, KY
Annco, Inc., Middleton, OH
Bethlehem Steel Bethlehem, PA
Bethlehem Steel, Bums Harbor, IN
Bethlehem Steel, Lackawanna, NY
Bethlehem Steel Sparrows Point, MD
Geneva Steel Orem, UT
Gulf States Steel Gadsden, AL
Inland Steel East Chicago, IN
LTV Steel Pittsburgh, PA
LTV Steel Chicago, IL
LTV Steel Cleveland, OH
LTV Steel, Warren, OH
National Steel Granite City, IL
National Steel Ecorse, MI
USS, Div. of USX Corp., Claiflon, PA
USS, Div. of USX Corp., Gary, IN
Wheeling-Pittsburgh Steel East SteubenviUe,
WV
Total
No. of
batteries
2
•
2
3
3
2
2
3
1
2
6
5
1
2
1
2
1
12
6
4
58
Total No. of
ovens
100
146
203
284
164
152
210
208
130
446
315
60
126
85
90
78
816
422
224
4,259
Total capacity, Mg/d
(ton/d)
1,450 (1,600)
2,450 (2,700)
4,130 (4,540)
3,580 (3,940)
3,980 (4,380)
1,700 (1,870)
3,700 (4,070)
2,050 (2,250)
2,550 (2,800)
5,250 (5,780)
4,910 (5,400)
1,450 (1,600)
2,910 (3,200)
1,360 (1,500)
1,380 (1,520)
840 (925)
11,490 (12,640)
6,490 (7,140)
3,450 (3,800)
65,120 (71,660)
4 Huskanen, 1991.
June 1996
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Figure 4-13
Coke Oven Locations
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 ail.; 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
June 1996
4-67
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magnitude of this source category's emissions. Emission factors used in Germany for coke production
range from 0.01 to O.OSg 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).
4.2.14 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 1992, there were 32 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 110 different locations. In addition, there are 72 companies
with distillation capacities of less than 100,000 barrels per calendar day (National Petroleum Refiners
Association, 1992).
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 processes to rearrange the structures of molecules; and polymerization and alkylation to
combine small molecules into larger ones (U.S. EPA, 1993a).
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, 1993a).
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, 1993a). 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, 1993a). 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.
June 1996 4-68 SAB REVIEW DRAFT
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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 4-25 presents mercury emissions from these miscellaneous sources.
4.3.1 Geothennal 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 1992. For dry-steam plants, steam is pumped from geothermal reservoirs
to turbines at a temperature of about 180°C (360°F) and a pressure of 7.9 bars absolute (U.S. EPA,
1993a). For water-dominated plants, water exists in the producing strata at a temperature of
approximately 270°C (520°F) and at a pressure slightly higher than hydrostatic (U.S. EPA, 1993a). 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). Table 4-26 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, 1993a).
Mercury emissions at geothermal power plants are documented to result from two sources:
off-gas ejectors and cooling towers. Table 4-27 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, 1993a). If
significant process modifications or changes hi control strategies have been incorporated since 1977,
the emission factors reported in Table 4-27 may no longer be valid.
Multiplying the emission factors in Table 4-27 by the total capacity shown in Table 4-26
(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 1992. Because the emission factors used to
generate this estimate have limited reliability, this emission estimate has a high degree of uncertainty.
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 cancelled 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, 1993a). Commercial mercury use in explosives ceased prior to 1970 (U.S.
EPA, 1992a).
June 1996 4-69 SAB REVIEW DRAFT
-------
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Table 4-26
1992 U.S. Geothermal Power Plants8
FacUity
The Geysers, CA
Salton Sea, CA
Heber, CA
East Mesa, CA
Coso, CA
Casa Diablo, CA
Amedee, CA
Wendel, CA
Dixie Valley, NV
Steamboat Hot Springs, NV
Beowawe Hot Springs, NV
Desert Peak, NV
Wabuska Hot Springs, NV
Soda Lake, NV
StiUwater, NV
Empire and San Emidio, NV
Roosevelt Hot Springs, UT
Cove Fort, UT
Type
Dry-steam
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Water-dominated
Net Capacity (MW)
1,805.7
218.3
47.0
106.0
247.5
34.0
2.0
0.7
57.0 '
19.3
16.7
9.0
1.7
15.7
12.5
3.2
20.0
12.1
Total 2,628.4
• Marshal, 1993.
Table 4-27
Mercury Emission Factors for Geothermal Power Plants3
Source
Off-gas ejectors
Cooling tower exhaust
Emission factor range
g/MWe/hr
0.00075 - 0.02
0.026 - 0.072
Average emission factor
g/MWe/hr
0.00725
0.05
Ib/MWe/hr
0.00002
0.0001
* Robertson et al, 1977.
June 1996
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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 hi 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 modei 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 peer review
panel that reviewed this work 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 0.86 g
mercury per person per year. Based on data submitted during the 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).
The similarity between the U.S. inventory and other countries where coal is also the major source of
energy lends credibility to the nationwide estimate presented in this report for the U.S.
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 220 Mg (243 tons) of annual mercury emission into the atmosphere by
anthropogenic sources: combustion point sources currently account for 85 percent, manufacturing
point sources contribute 13 percent, area sources contribute 1 percent, and miscellaneous point sources
contribute 1 percent Four specific source categories account for approximately 83 percent of the total
anthropogenic emissions - medical waste incineration (27 percent),
June 1996 5-1 SAB REVIEW DRAFT
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Table 5-1
Best Point Estimates of National Mercury Emission Rates by Category
Source of mercury1
Area sources
Lamp breakage
General lab use
Dental prep and use
Mobile sources
Paint use
Agricultural burning
Landfills
Point sources
Combustion sources
MWIs*
MWCs
Utility boilers
Coal
Oil
Natural gas
Commercial/industrial boilers
Coal
Oil
Residential boilers
Coal
Oil
SSIs
Crematories
Wood-fired boilers11
Hazardous waste combustors1
Manufacturing sources
Primary lead
Secondary Hg production
Chlor-alkali
Portland cement
Primary copper1
Lime manufacturing
Electrical apparatus
Instruments
Carbon black
Fluorescent lamp recycling
Batteries
Primary Hg production
Mercury compounds
Byproduct coke
Refineries
Miscellaneous sources
Geothermal power
Turf products
Pigments, oil, etc.
TOTAL
1990-1993
Mg/yrb>c
2.8
1.4
0.7
0.7
d
e
d
d
217.3
186.9
58.8
\ 50
46.5
(46.3)8
(0.23)
(0.002)
263
(20.7)
(5.5)
3.2
(0.5)
(2.7)
1.7
0.4
03
d
29.1
8.2
6.7
5.9
5.9
0.6
0.6
0.42
0.5
0.23
0.005
0.02
d
d
d
d'
1.3
1.3
e
e
220.1
1990-1993
tons/yrb'c
3.1
1.5
0.8
0.8
d
e
d
d
239.4
205.9
64.7
55
513
- (51)
(0.25)
(0.002)
29
(22.8)
(6.0)
3.5
(0.6)
(3.0)
1.8
0.4
0.3
d
32
9.0
7.4
6.5
6.5
0.7
0.7
0.46
0.5
0.25
0.006
0.02
d
d
d
d
1,4
1.4
e
e
242.5
% of Total
Inventory
13
0.6
03
03
d
e
d
d
98.7
84.9
26.7
22.7
21.2
(21.0)
(0.1)
(0.0)
12.0
(9.4)
(2.5)
1.4
0.2
(1.2)
0.7
0.2
0.1
d
13.2
3.7
3.1
2.7
2.7
0.3
03
0.2
0.2
0.1
0.002
0.0
d
d
d
d
0.6
0.6
e
e
100.0
* MWC » Municipal waste combastor, MWI 3 medical waste incinerator, SSI a sewage sludge incinerator.
* Numbers do not add exactly because of rounding.
° Where available, emissions estimates for 1995 ate discussed in the text However, these 199S estimates were not used in any of the modeling
analyses.
d Insufficient information to estimate 1990 emissions.
° Mercury has been phased out of use.
f In the course of an MWI rulemaking, with the receipt of new data, U.S. EPA expects to revise the mercury emission estimate for MWIs downward.
* Parentheses denote subtotal within a larger point source category.
h Includes boilers only; does not include residential wood combustion (wood stoves).
1 In 1995 incinerators and lightweight aggregate labs (not cement kilns) were estimated to emit 5.0 tons of mercury.
1 1990 emissions are estimated for only one source, which ceased operations in February 199S. The nationwide estimate for 199S is O.OS tons.
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municipal waste combustion (23 percent), utility boilers (21 percent), and commercial/industrial boilers
(12 percent).
Based on this information, six source categories were selected for the local impact analyses in
Volume HI of this report and the control technology assessment described in Volume VII 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, chlor-alkali plants,
primary copper smelters and primary lead smelters. 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 hi detail in Appendix F to Volume III of this report and for
the control technology assessment, in Appendix B of Volume VII 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 hi 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.
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.
Throughout development of the inventory, an attempt was made to incorporate, to the extent
possible, the most recent emissions data for each of the source categories. Not all of the changes
could be incorporated into the subsequent modeling analyses. As a result, there are three discrepancies
between the inventory described in this document and the emissions modeled in the long-range
transport analysis. First in this document aggregate emissions for the chlor-alkali industry are
estimated to be 5.9 Mg/year (6.5 tons/year) based on 1991 industry surveys. By comparison, the long-
range transport analysis modeled 6.5 Mg/year (7.2 tons/year) based on earlier data submitted to the
Toxic Release Inventory. The second discrepancy is for mercury emissions from paint application.
The mercury inventory was updated to reflect that no mercury has been added to paint since 1991.
The long-range transport analysis uses a previous estimate of 4 Mg/year (4.4 tons/year) to account for
previously applied paint that may still be out-gassing mercury (U.S. EPA, 1993d). The third
discrepancy is that the inventory was revised to reflect 0.23 Mg/year (0.25 tons/yr) of emissions from
oil-fired utility boilers. This revision was based on a re-analysis of the detection limits for residual oil
samples. The long-range transport analysis uses an earlier estimate of 3.25 tons/yr for this source
category. None of these discrepancies is believed to be significant enough to affect the mercury
deposition patterns predicted by the long-range transport model.
June 1996 5-4 SAB REVIEW DRAFT
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Table 5-3
Mercury Area Sources Allocation Methodology
Area Source Category
Latex Paint Application
Mercury Lamp Breakage
General Laboratory Usage
Dental Preparation
Residential Coal Combustion
Residential Oil Combustion
Industrial/Commercial
Boilers
Coal
Oil
Crematories
Medical Waste Incinerators
Emissions
Mg/year
(tons/yr)
4 (4.4)
1.4 (1.5)
0.7 (0.8^
6.7 (0.8)
0.5 (0.6)
2.7 (3.0)
20.7 (22.8)
5.5 (6.0)
0.4 (0.4)
58.8 (64.7)
Allocation Method
•
Nationwide estimate allocated on a per
capita basis (1980 Census).
Nationwide estimate allocated to counties
on a per capita basis (1980 Census).
Nationwide estimate allocated on a per
capita basis (1980 Census data).
Nationwide estimate allocated by number
of dental establishments. SIC 8020,
8072.
Nationwide estimate allocated to counties
that reported residential coal combustion
in the 1980 Census. Apportionment to
counties on a per capita basis.
Nationwide estimate allocated by State
based on fuel consumption (U.S.
Department of Energy, 1992).
Apportionment to counties within State
on a per capita basis.
Nationwide estimates allocated by State
based on fuel consumption (U.S.
Department of Energy, 1992).
Apportionment to counties within State
on a per capita basis.
Statewide emissions estimate allocated to
counties on a per capita basis.
Nationwide estimate allocated by the
number of hospital beds per county.
SIC = Standard Industrial Classification
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In addition, it should be noted that the long-range transport analysis utilized the emission
estimates for 1990 from this inventory. For some source categories, estimates for 1995 are given
where new information became available prior to publication. However, it is the 1990 estimates that
were modeled.
Trends in Mercury Emissions
It is difficult to predict with certainty 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 mercury is a trace contaminant of the fuel.
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.
There has been a real success in the U.S. in the dramatic drop in mercury emissions from
manufacturing over the past decade. Current emissions of mercury from manufacturing sources are
generally low (with the exception of chlor-alkali plants using the mercury cell process). The emissions
of mercury are more likely to occur when the product is broken or discarded. Therefore, in terms of
emission trends, one would expect that if the future consumption of mercury remains consistent with
the 1993 consumption rate, emissions from most manufacturing sources would remain about the same.
For industrial or manufacturing sources that use mercury hi products or processes, the overall
consumption of mercury is generally declining. Industrial consumption of mercury has declined by
about two thirds between 1988 (1508 Mg) and 1993 (558 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, as described in Chapter 2 of this document,
remained about the same between 1988 and 1993.
Secondary production of mercury (i.e., recovering mercury from waste products) has increased
significantly over the past few years. Of the 558 Mg of mercury used in industrial processes in 1993,
63 percent was provided by secondary mercury producers. This represents 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., thermometers).
As a result there is potential for mercury emissions from this source category to increase.
The largest identifiable sources of mercury emissions currently are municipal waste combustors
and medical waste incinerators. Emissions from these source categories are expected to decline
significantly 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 4.1.1 and 4.1.2 of this document, the
U.S. EPA has final rules for MWCs and proposed rules for MWIs that will reduce mercury emissions
from both of these source categories by about 90 percent In addition to this federal action, a number
of states (including Minnesota, Florida and New Jersey) have implemented mandatory recycling
programs to reduce the 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.
After municipal solid waste and medical waste incinerators have been controlled, the largest
remaining identified source of mercury emissions will be fossil fuel combustion by utility boilers,
particularly coal combustion. Future trends in mercury emissions from this source category are largely
June 1996 5-7 SAB REVIEW DRAFT
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dependent on both the nation's future energy needs and the fuel chosen to meet those needs. Another
factor is what actions the utility industry may take in the future to meet air quality requirements under
the Clean Air Act
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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 dramatically over the past
decade.
• Prior to 1995, municipal waste combustors and medical waste incinerators were the largest
identifiable source of mercury emissions to the atmosphere. Regulations finalized for
municipal waste combustors and proposed for medical waste incinerators will reduce emissions
from these sources by 90 percent.
• After emissions from municipal solid waste combustors and medical waste incinerators have
been reduced, combustion of fossil fuels, particularly coal, will be the largest remaining source
of mercury emissions to the atmosphere.
• 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.
• Anthropogenic sources hi the United States emit approximately 220 Mg (243 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 1990-1993 timeframe. Emission
estimates for 1995 are about 40 tons lower.
• In the United States, land areas east of the Rocky Mountains have the highest concentration of
emissions from anthropogenic sources in the U.S.
• The land 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 land 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 hi 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.
June 1996 6-1 SAB REVIEW DRAFT
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• 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 that 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:
• Emissions test data for source categories where there is currently insufficient data to estimate
national emissions, including mobile sources, landfills, 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 which
could include these: secondary mercury production (i.e., recycling), commercial and industrial
boilers, electric lamp breakage, iron and steel manufacturing and primary lead smelting.
Based on trends in mercury use and emissions, the U.S. EPA predicts the following:
• A significant (90 percent) decrease will occur in mercury emissions from municipal waste
combustors and medical waste incinerators when the regulations put forth 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.
June 1996 6-2 SAB REVIEW DRAFT
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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,
landfills, 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, 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 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.
June 1966 7-1 SAB REVIEW DRAFT
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- \
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June 1996 8-11 SAB REVIEW DRAFT
-------
APPENDIX A
INFORMATION ON LOCATIONS OF AND EMISSIONS FROM
COMBUSTION SOURCES
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Table A-2
Estimates of Coal, Natural Gas, and Oil Consumption
in the Commercial/Industrial Sector Per State
(Trillion Btu)
State
Alabama
Alaska
Arizona
Arkansas
California
Colorado
Connecticut
Delaware
Dist of Col.
Florida
Georgia
Hawaii
Idaho
Illinois
Indiana
Iowa
Kansas
Kentucky
Louisiana
Maine
Maryland
Massachusetts
Michigan
Minnesota
Mississippi
Missouri
Montana
Coal
Bituminous coal
and lignite
144.6
5.1
133
5.8
65.1
16.2
0.0
5.8
1.1
30.1
56.5
0.7
9.6
154.5
350.6
56.2
3.8
89
16
6.0
58.1
23
122.2
25.9
6.3
34.5
4.7
Anthracite
0.4
0.0
0.0
0.0
0.0
0.0
0.1
0.0
0.0
0.1
a
0.0
0.0
0.1
0.1
0.7
0.0
0.5
0.0
0.1
a
0.4
a
0.0
0.0
0.0
0.0
Total
145
5.1
133
5.8
65.1
16.2
0.1
5.8
1.1
30.2
56.5
0.7
9.6
154.6
350.7
56.9
3.8
89.5
16
6.1
58.1
2.7
122.2
25.9
63
34.5
4.7
Natural gas
185.0
277.2
483
153.6
900.6
133.4
56.7
21.4
13.6
133.7
217.2
2.4
32.8
4863
300.8
135.2
213.8
107.8
1242.4
3.7
88.2
98.1
4683
167.0
129.8
115.1
24.5
Petroleum
Distillate and
residual
45.6
16.0
21.1
23.4
138.2
18.1
23
4.5
2.9
39.2
31.1
7.7
17.9
53.4
32.7
27.0
24.4
34.2
84.8
13.9
22.3
49.0
29.9
33.0
36.4
23.6
16.9
June 1996
A-4
SAB REVIEW DRAFT
-------
Table A-2 (continued)
Estimates of Coal, Natural Gas, and Oil Consumption
in the Commercial/Industrial Sector Per State
(Trillion Btu)
State
Nebraska
Nevada
New Hampshire
New Jersey
,New Mexico
New York
North Carolina
North Dakota
Ohio
Oklahoma
Oregon
Pennsylvania
Rhode Island
South Carolina
South Dakota
Tennessee
Texas
Utah
Vermont
Virginia
Washington
West Virginia
Wisconsin
Wyoming
United States
Coal
Bituminous coal
and lignite
4.6
3.9
0.5
6.8
1.0
84.0
77.1
87.5
258.0
12.7
1.4
3823
0.0
57.9
3.9
100.0
61.7
52.7
0.0
121.4
6.1
127.0
47.2
42.9
2824.6
Anthracite
a
a
0.3
03
0.0
2.1
0.0
0.0
0.3
0.0
0.0
15.1
0.1
0.2
0.0
0.3
a
0.0
0.1
0.3
0.0
0.1
0.1
0.0
21.8
Total
4.6
3.9
0.8
7.1
1
86.1
77.1
87.5
2583
12.7
1.4
397.4
0.1
58.1
3.9
100.3
61.7
52.7
0.1
121.7
6.1
127.1
473
42.9
2842.5
Natural gas
613
23.2
8.4
211.2
115.0
305.7
121.2
223
444.7
350.7
71.0
393.5
83
105.1
14.7
158.7
2373.3
77.8
3.9
121.1
120.7
84.6
189.4
83.1
11226.0
Petroleum
Distillate and
residual
25.5
21.0
9.4
57.2
16.4
95.6
28.3
16.0
39.5
21.1
24.4
69.3
5.3
14.9
13.1
20.7
169.9
10.9
6
31.6
38.2
18.2
31.6
14.5
1668.8
a Number less than 0.05
Source: U.S. Department of Energy. State Energy Data Report Report No. DOE/EIA-0214(40). May 1992.
June 1996
A-5
SAB REVIEW DRAFT
-------
Table A-3
Estimates of Mercury Emissions From Coal-Fired
Commercial/Industrial Boilers on a Per-State Basis For 1991
State
Alabama
Alaska
Arizona
Arkansas
California
Colorado
Connecticut
Delaware
Dist of Col.
Florida
Georgia
Hawaii
Idaho
Illinois
Indiana
Iowa
Kansas
Kentucky
Louisiana
Maine
Maryland
Massachusetts
Michigan
Minnesota
Mississippi
Missouri
Montana
Nebraska
Nevada
New Hampshire
New Jersey
New Mexico
New York
North Carolina
North Dakota
Ohio
Oklahoma
Oregon
Pennsylvania
Rhode Island
South Carolina
South Dakota
Tennessee
Texas
Utah
Vermont
Virginia
Washington
West Virginia
Wisconsin
Wyoming
United States
Coal consumption, trillion Btu
Bituminous coal and
lignite
144.6
5.1
13.3
5.8
65.1
16.2
0.0
5.8
1.1
30.1
56.5
0.7
9.6
154.5
350.6
56.2
3.8
89
16
6.0
58.1
2.3
122.2
25.9
6.3
34.5
4.7
4.6
3.9
0.5
6.8
1.0
84.0
77.1
87.5
258.0
12.7
1.4
382.3
0.0
57.9
3.9
100.0
61.7
52.7
0.0
121.4
6.1
127.0
47.2
42.9
2824.6
Anthracite
0.4
0.0
0.0
0.0
0.0
0.0
0.1
0.0
0.0
0.1
b
0.0
0.0
0.1
0.1
0.7
0.0
0.5
0.0
0.1
b
0.4
b
0.0
0.0
0.0
0.0
b
b
0.3
0.3
0.0
2.1
0.0
0.0
0.3
0.0
0.0
15.1
0.1
0.2
0.0
0.3
b
0.0
0.1
0.3
0.0
0.1
0.1
0.0
21.8
Total
145
5.1
13.3
5.8
65.1
16.2
0.1
5.8
1.1
30.2
56.5
0.7
9.6
154.6
350.7
56.9
3.8
89.5
16
6.1
58.1
2.7
12Z2
25.9
6.3
34.5
4.7
4.6
3.9
0.8
7.1
1
86.1
77.1
87.5
258.3
12.7
1.4
397.4
0.1
58.1
3.9
100.3
61.7
52.7
0.1
121.7
6.1
127.1
47.3
42.9
2842.5
Mercury emissions2
Ton/Yrc
1.2
0.0
0.1
0.0
0.5
0.1
0.0
0.0
0.0
0.2
0.5
0.0
0.1
1.2
2.8
0.5
0.0
0.7
0.1
0.0
0.5
0.0
1.0
0.2
0.1
0.3
0.0
0.0
0.0
0.0
0.1
0.0
0.7
0.6
0.7
2.1
0.1
0.0
3.2
0.0
0.5
0.0
0.8
0.5
0.4
0.0
1.0
0.0
1.0
0.4
0.3
22.8
Mg/Yr
1.1
0.0
0.1
0.0
0.5
0.1
0.0
0.0
0.0
0.2
0.4
0.0
0.1
1.1
2.6
0.4
0.0
0.7
0.1
0.0
0.4
0.0
0.9
0.2
0.0
0.3
0.0
0.0
0.0
0.0
0.1
0.0
0.6
0.6
0.6
1.9
0.1
0.0
2.9
0.0
0.4
0.0
0.7
0.4
0.4
0.0
0.9
0.0
0.9
0.3
0.3
20.7
* Mercury emission factors of 16 Ib Kg/trillion Btu and 18 Ib Hg/trillion Btu were used for bituminous and anthracite coal, respectively. No
control of emissions from commercial/industrial boilers was assumed.
b Number less than 0.05.
c Emissions less than 100 pounds/year for an entire State are reported as zero.
June 1996
A-6
SAB REVIEW DRAFT
-------
Table A-4
Estimates of Mercury Emissions From Oil-Fired
Commercial/Industrial Boilers On a Per-State Basis For 1991
State
Alabama
Alaska
Arizona
Arkansas
California
Colorado
Connecticut
Delaware
Dist. of Col.
Florida
Georgia
Hawaii
Idaho
Illinois
Indiana
Iowa
Kansas
Kentucky
Louisiana
Maine
Maryland
Massachusetts
Michigan
Minnesota
Mississippi
Missouri
Montana
Nebraska
Nevada
New Hampshire
New Jersey
New Mexico
New York
North Carolina
North Dakota
Ohio
Oklahoma
Oregon
Pennsylvania
Rhode Island
South Carolina
South Dakota
Tennessee
Texas
Utah
Vermont
Virginia
Washington
West Virginia
Wisconsin
Wyoming
United States
Petroleum consumption, trillion
Btu
Distillate and residual
45.6
16.0
21.1
23.4
138.2
18.1
23
4.5
2.9
39.2
31.1
7.7
17.9
53.4
32.7
27.0
24.4
34.2
84.8
13.9
22.3
49.0
29.9
33.0
36.4
23.6
16.9
25.5
21.0
9.4
57.2
16.4
95.6
28.3
16.0
39.5
21.1
24.4
69.3
5.3
14.9
13.1
20.7
169.9
10.9
6
31.6
38.2
18.2
31.6
14.5
1,668.8
Mercury emissions9
Ton/Yr
0.16
0.06
0.08
0.08
0.50
0.07
0.08
0.02
0.01
0.14
0.11
0.03
0.06
0.19
0.12
0.10
0.09
0.12
0.31
0.05
0.08
0.18
0.11
0.12
0.13
0.08
0.06
0.09
0.08
0.03
0.21
0.06
0.34
0.10
0.06
0.14
0.08
0.09
0.25
0.02
0.05
0.05
0.07
0.61
0.04
0.02
0.11
0.14
0.07
0.11
0.05
6.01
Mg/Yr
0.15
0.05
0.07
0.08
0.45
0.06
0.08
0.01
0.01
0.13
0.10
0.03
0.06
0.17
0.11
0.09
0.08
0.11
0.28
0.05
0.07
0.16
0.10
0.11
0.12
0.08
0.06
0.08
0.07
0.03
0.19
0.05
0.31
0.09
0.05
0.13
0.07
0.08
0.23
0.02
0.05
0.04
0.07
0.56
0.04
0.02
0.10
0.13
0.06
0.10
0.05
5.46
a Mercury emission factor for distillate oil is 7.2 Ib Hg/tnllion Btu. Calculation was performed assuming that all pollution control devices
provide no mercury reduction.
s Number less than 0.05.
June 1996
A-7
SAB REVIEW DRAFT
-------
Table A-5
Estimates of Coal, Natural Gas, and Oil Consumption
in the Residential Sector Per State (Trillion Btu)
State
Alabama
Alaska
Arizona
Arkansas
California
Colorado
Connecticut
Delaware
Distof CoL
Florida
Georgia
Hawaii
Idaho
Illinois
Indiana
Iowa
Kyn.Bay
Kentucky
Louisiana
Maine
Maryland
Massachusetts
Michigan
Minnesota
Mississippi
Missouri
Montana
Nebraska
Nevada
New Hampshire
New Jersey
New Mexico
New York
North Carolina
North Dakota
Ohio
Oklahoma
Oregon
Pennsylvania
Rhode Island
South Carolina
South Dakota
Tennessee
Texas
Utah
Vermont
Virginia
Washington
West Virginia
Wisconsin
Wyoming
United States
Coal
Bituminous coal and
lignite
0.9
2.7
a
a
0.2
0.4
0.0
0.2
0.6
a
0.2
0.0
0.5
2.1
4.3
2.0
a
1.3
0
0.3
0.4
0.3
2.3
1.1
0.0
22
0.4
a
a
0.0
0.0
a
1.2
1.4
0.7
5.5
0.0
a
2.9
0.0
0.1
a
1.8
0.1
2.2
0.0
2.1
0.5
1.6
a
0.9
43.4
Anthracite
0.0
0.0
0.0
0.0
0.0
0.0
0.2
0.0
0.0
0.0
0.0
0.0
0.0
a
a
a
0.0
0.0
0.0
0.2
0.1
0.4
0.0
0.0
0.0
0.0
0.0
0.0
a
02
0.2
0.0
2.0
0.0
0.0
a
0.0
0.0
14.7
0.1
0.0
0.0
0.1
0.0
0.0
0.1
a
0.0
0.0
a
0.0
18.3
Total
0.9
2.7
0
0
0.2
0.4
0.2
0.2
0.6
0
0.2
0
OS
2.1
4.3
2
0
1.3
0
0.5
0.5
0.7
2.3
1.1
0
22
0.4
0
a
0.2
0.2
0
3.2
1.4
0.7
5.5
0
0
17.6
0.1
0.1
0
1.9
0.1
2.2
0.1
2.1
0.5
1.6
0
0.9
61.7
Natural gas
46.7
13.4
29.3
39.5
530.8
92.4
38.7
7.4
15.3
14.1
92.7
0.6
8.8
451.9
143.1
71.9
71.3
58.5
55.6
0.7
68.2
110.5
342.2
107.4
25.9
117.2
17.3
40.8
17.7
6.0
176.0
29.7
347.8
36.1
9.5
321.0
65.9
23.9
248.9
18.2
18.9
10.4
48.0
220.8
47.3
2.1
53.6
41.6
34.9
114.7
12.6
4.518.8
Petroleum distillate
and residual
0.1
10.2
0.1
a
1.3
0.2
66.6
5.6
0.9
1.4
1.5
a
3.1
7.0
10.0
4.6
0.1
3.8
0.1
29.3
25.0
100.7
24.3
18.8
a
2.1
1.7
1.0
1.4
19.8
67.0
0.1
154.5
20.7
4.9
23.8
a
10.4
99.1
14.9
5.9
4.7
1.4
a
0.8
11.2
29.8
17.5
3.3
27.0
0.1
837.8
* Number less than 0.05.
Source: U.S. Department of Energy. State Energy Data Report Report No. DOE/EIA-0214(40). May 1992.
June 1996
A-8
SAB REVIEW DRAFT
-------
Table A-6
Estimates of Mercury Emissions From
Coal-Fired Residential Boilers on a Per-State Basis For 1991
State
Alabama
Alaska
Arizona
Arkansas
California
Colorado
Connecticut
Delaware
Dist of CoL
Florida
Georgia ,
Hawaii
Idaho
Illinois
Tndjiffla
Iowa
Kansas
Kentucky
Louisiana
Maine
Maryland
Massachusetts
Michigan
Minnesota
Mississippi
Missouri
Montana
Nebraska
Nevada
New Hampshire
New Jersey
New Mexico
New York
North Carolina
North Dakota
Ohio
Oklahoma
Oregon
Pennsylvania
Rhode Island
South Carolina
South Dakota
Tennessee
Texas
Utah
Vermont
Virginia
Washington
West Virginia
Wisconsin
Wyoming
United States
Coal consumption, trillion Btu
Bituminous coal and
lignite
0.9
2.7
b
b
02
0.4
0.0
0.2
0.6
b
0.2
0.0
0.5
2.1
4.3
2.0
b
1.3
0
0.3
0.4
0.3
2.3
1.1
0.0
22
0.4
b
b
0.0
0.0
b
1.2
1.4
0.7
5J
0.0
b
2.9
0.0
0.1
b
1.8
0.1
2.2
0.0
2.1
0.5
1.6
b
0.9
43.4
Anthracite
0.0
0.0
0.0
0.0
0.0
0.0
0.2
0.0
0.0
0.0
0.0
0.0
0.0
b
b
b
0.0
0.0
0.0
0.2
0.1
0.4
0.0
0.0
0.0
0.0
0.0
0.0
b
0.2
0.2
0.0
2.0
0.0
0.0
b
0.0
0.0
14.7
0.1
0.0
0.0
0.1
0.0
0.0
0.1
b
0.0
0.0
b
0.0
18.3
Total
0.9
2.7
0
0
0.2
0.4
0.2
0.2
0.6
0
0.2
0
0.5
2.1
4.3
2
0
1.3
0
0.5
0.5
0.7
2.3
1.1
0
22
0.4
0
b
0.2
0.2
0
3.2
1.4
0.7
5.5
0
0
17.6
0.1
0.1
0
1.9
0.1
2.2
0.1
2.1
0.5
1.6
0
0.9
61.7
Mercury emissions1
Ton/Yr
0.007
0.022
0.000
0.000
0.002
0.003
0.002
0.002
0.005
0.000
0.002
0.000
0.004
0.017
0.034
0.016
0.000
0.010
0.000
0.004
0.004
0.006
0.018
0.009
0.000
0.018
0.003
0.000
0.000
0.002
0.002
0.000
0.028
0.011
0.006
0.044
0.000
0.000
0.156
0.001
0.001
0.000
0.015
0.001
0.018
0,001
0.017
0.004
0.013
0.000
0.007
0.512
Mg/Yr
0.007
0.020
0.000
0.000
0.001
0.003
0.002
0.001
0.004
0.000
0.001
0.000
0.004
0.015
0.031
0.015
0.000
0.009
0.000
0.004
0.004
0.005
0.017
0.008
0.000
0.016
0.003
0.000
0.000
0.002
0.002
0.000
0.025
0.010
0.005
0.040
0.000
0.000
0.141
0.001
0.001
0.000
0.014
0.001
0.016
0.001
0.015
0.004
0.012
0.000
0.007
0.465
* Mercury emission factors of 16 Ib Hg/trillion Btu and 18 Ib Hg/trillion Btu were used for bituminous and anthracite coal, respectively. No
control of emissions from residential boilers was assumed.
b Number less than 0.05.
June 1996
A-9
SAB REVIEW DRAFT
-------
Table A-7
Estimates of Mercury Emissions From Oil-Fired
Residential Boilers on a Per-State Basis For 1991
State
Alabama
Alaska
Arizona
Arkansas
California
Colorado
Connecticut
Delaware
Dist of CoL
Florida
Georgia
Hawaii
Idaho
Illinois
Indiana
Iowa
Kansas
Kentucky
Louisiana
Maine
Maryland
Massachusetts
Michigan
Minnesota
Mississippi
Missouri
Montana
Nebraska
Nevada
New Hampshire
New Jersey
New Mexico
New York
North Carolina
North Dakota
Ohio
Oklahoma
Oregon
Pennsylvania
Rhode Island
South Carolina
South Dakota
Tennessee
Texas
Utah
Vermont
Virginia
Washington
West Virginia
Wisconsin
Wyoming
United States
Petroleum consumption, trillion Btu
Distillate and residual
0.1
10.2
0.1
*
1.3
0.2
66.6
5.6
0.9
1.4
1.5
*
% 3.1
7.0
10.0
4.6
0.1
3.8
0.1
29.3
25.0
100.7
24.3
18.8
«
2.1
1.7
1.0
1.4
19.8
67.0
0.1
154.5
20.7
4.9
23.8
*
10.4
99.1
14.9
5.9
4.7
1.4
*
0.8
11.2
29.8
17.5
3.3
27.0
0.1
837.8
Mercury emissions'
Ton/Yr
0.0004
0.0367
0.0004
0.0000
0.0047
0.0007
0.2398
0.0202
0.0032
0.0050
0.0054
0.0000
0.0112
0.0252
0.0360
0.0166
0.0004
0.0137
0.0004
0.1055
0.0900
0.3625
0.0875
0.0677
0.0000
0.0076
0.0061
0.0036
0.0050
0.0713
0.2412
0.0004
0.5562
0.0745
0.0176
0.0857
0.0000
0.0374
0.3568
0.0536
0.0212
0.0169
0.0050
0.0000
0.0020
0.0403
0.1073
0.0630
0.0119
0.0972
0.0004
3.02
Mg/Yr
0.0003
0.0334
0.0003
0.0000
0.0043
0.0007
0.2180
0.0183
0.0029
0.0046
0.0049
0.0000
0.0101 \
0.0229
0.0327
0.0151
0.0003
0.0124
0.0003
0.0959
0.0818
0.3296
0.0795
0.0615
0.0000
0.0069
0.0056
0.0033
0.0046
0.0648
0.2193
0.0003
0.5056
0.0677
0.0160
0.0779
0.0000
0.0340
0.3243
0.0488
0.0193
0.0154
0.0046
0.0000
0.0026
0.0367
0.0975
0.0573
0.0108
0.0884
0.0003
2.74
* Mercury emission factor for Hi«tjflat«< oil is 7.2 Ib Kg/trillion Btu. Calculations performed under the assumption that air pollution control
devices provide no mercury reduction.
b Number less than 0.05.
June 1996
A-10
SAB REVIEW DRAFT
-------
Table A-8
1991 U.S. Crematory Locations by State*'1'
State
Alabama
Alaska
Arizona
Connecticut
Delaware
District of
Columbia
Florida
Georgia
Hawaii
Idaho
Illinois
Indiana
Iowa
Kansas
Kentucky
Louisiana
Maine
Maryland
Massachusetts
Michigan
Minnesota
Missouri
No. of
crematories
6
7
26
10
4
1
95
14
10
12
44
21
15
10
5
6
4
17
13
38
18
19
No. of
cremations0
1,138
790
10,189
4,260
1,165
d
46,775
2,684
3,495
1,949
12,083
3,636
2,241
1,559
1,192
1,853
2,656
5,587
8,104
13,431
5,662
4,637
State
Montana
Nebraska
Nevada
New York
North Carolina
North Dakota
Ohio
Oklahoma
Oregon
Pennsylvania
Rhode Island
South Carolina
South Dakota
Tennessee
Texas
Utah
Vermont
Virginia
Washington
West Virginia
Wisconsin
No. of
crematories
12
6
11
40
24
1
41
9
34
44
5
10
4
8
36
5
5
25
46
6
29
No. of
cremations
2,502
1,139
5,009
23,946
4,749
d
12,552
1,372
9,020
12,153
1,842
1,764
d
1,712
9,340
769
1,570
6,097
" 15,673
582
5,541
a Vander Most and Veldt, 1992.
b Does not include the number of cremations in the District of Columbia, Arkansas, California, New Hampshire, New
Mexico, Colorado, North Dakota, South Dakota, and Wyoming, or other U.S. territories.
c 1990 data; 1991 data unavailable.
d No information available.
June 1996
A-ll
SAB REVIEW DRAFT
-------
Table A-9
Existing MWC Facilities (As of December, 1991)
Facility
Parsons (SOfflO)a
Juneau
Kyparuk (ARCO)a
Pradhoe Bay*
Shemya (Air Force Base)*
Sitka (Sheldon Jackson
College)
Huntsville
Tuscaloosaa
Augusta*
Batesville
Blythevffle
Kensett*
North Little Rock*
Osceola
Stuttgart
Los Angeles County
Long Beach (SERRF)
Stanislaus County
Bridgeport
Bristol
MID-Connecticut
New Cannan*
Southeastern
Stamford H*
Stamford I*
Wallingford
City
Endicott
Juneau
Kyparuk
Pradhoe Bay
Shemya
Sitka
Huntsville
Tuscaloosa
Augusta
Batesville
BlytheviUe
Kensett
North Little Rock
Osceola
Stuttgart
Commerce
Long Beach
Modesto
Bridgeport
Bristol
Hartford
New Cannan
Preston
Stamford
Stamford
Wallingford
State
AK
AK
AK
AK
AK
AK
AL
AL
AR
AR
AR
AR
AR
AR
AR
CA
CA
CA
CT
CT
CT
CT
CT
CT
CT
CT
Capacity
tons/yr
4,380
25,550
4,380
36,500
7,300
9,125
251,850
109,500
7,300
36,500
25,550
5,475
36,500
18,250
21,900
138,700
503,700
' 292,000
821,250
237,250
730,000
45,625
219,000
131,400
54,750
153,300
Mg/yr
3,982
23,227
3,982
33,182
6,636
8,295
228,955
99,545
6,636
33,182
23,227
4,977
33,182
16,591
19,909
126,091
457,909
265,455
746,591
215,682
663,636
41,477
199,091
119,455
49,773
139,364
a No longer operational.
June 1996
A-12
SAB REVIEW DRAFT
-------
Table A-9 (continued)
Existing MWC Facilities (as of December, 1991)
a No longer operational.
June 1996
Facility
Windham*
Washington2
Wilmington (Newcastle)*
Hillsborough County
FortMeadea
Broward County (South)
Pasco County
Monroe County
Lakeland*
MayportNAS
Dade County
Miami (Airport)
Lake County
Bay County
Broward County (North)
Pinellas County
McKay Bay
Palm Beach County
Savannah
Honolulu
Honolulu (Waipahu)*
Ames*
Burley (Cassia County)
Chicago NW
Indianapolis RRF
Louisville*
Springfield RRF
City
Windham
Washington
Wilmington
Brandon
FortMeade
Ft. Lauderdale
%
Hudson
Key West
Lakeland
Mayport
Miami
Miami
Okahumpka
Panama City
Pompano Beach
SL Petersburg
Tampa
West Palm Beach
Savannah
Honolulu
Honolulu
Ames
Burley
Chicago
Indianapolis
Louisville
Agawam
State
CT
DC
DE
FL
FL
FL
FL
• FL
FL
FL
FL
FL
FL
FL
FL
FL
FL
FL
GA
HI
HI
LA
ED
IL
IN
KY
MA
Capacity
tons/yr
39,420
365,000
219,000
438,000
9,490
821,250
383,250
54,750
109,500
17,520
1,095,000
21,900
192,720
186,150
821,250
1,095,000
365,000
730,000
182,500
788,400
219,000
73,000
18,250
584,000
862,130
365,000
131,400
Mg/yr
. 35,836
331,818
199,091
398,182
8,627
746,591
348,409
49,773
99,545
15,927
995,455
19,909
175,200
169,227
746,591
995,455
331,818
663,636
165,909
716,727
199,091
66,364
16,591
530,909
783,755
331,818
119,455
A-13
SAB REVIEW DRAFT
-------
Table A-9 (continued)
Existing MWC Facilities (as of December, 1991)
a No longer operational.
June 1996
Facility
Fall River*
Fnunififfhsun*
Haverhffl
Lawrence
Millbury
North Andover RESCO
Pittsfield RRF X
Rochester (SEMASS)
Saugus RESCO
Springfield*
Harford County
Baltimore (Pulaski)
Baltimore (RESCO)
Biddeford
Aroostook County*
Harpswell
Penobscot (Orrington)
Portland
Grosse Point Clinton
Greater Detroit Resource
Recovery Authority
Fisher Guide Division*
Kent County District Waste to
Energy Facility
Jackson County
SE Oakland County*
Central Wayne County
Sanitation Authority
City
Fall River
Framingham
Haverhill
Lawrence
Millbury
North Andover
Pittsfield
Rochester '
Saugus
Springfield
Aberdeen
Baltimore
Baltimore
Biddeford
Frenchville
Harpswell
Orrington
Portland
Clinton Township
Detroit
Detroit
Grand Rapids
Jackson
Madison Heights
Inkster
State
MA
MA
MA
MA
MA
MA
MA
MA
MA
MA
MD
MD
MD
ME
ME
ME
ME
ME
MI
MI
MI
MI
MI
MI
MI
Capacity
tons/yr
. 219,000
182,500
602,250
259,150
547,500
547,500
131,400
657,000
547,500
131,400
131,400
438,000
821,250
219,000
18,250
5,110
262,800
182,500
219,000
1,204,500
36,500
228,125
73,000
219,000
36,500
Mg/yr
199,091
165,909
547,500
235,591
497,727
497,727
119,455
597,273 -
497,727
119,455
119,455
398,182
746,591
199,091
16,591
4,645
238,909
165,909
199,091
1,095,000
33,182
207,386
66,364
199,091
16,590
A-14
SAB REVIEW DRAFT
-------
Table A-9 (continued)
Existing MWC Facilities (as of December, 1991)
Facility
Alexandria
Duluth
Anoka County (Elk River)
Fergus Falls
Polk County
Mankato
Hennepin County
Perham
Red Wing
Red Wing
Olmstead County
Savage
Thief River Falls (Hospital)
Thief River Falls (Hospital)
Ft Leonard Wood*
Pascagoula
Livingston (Park County)
Mecklenburg Co.
Cherry Point Marine St.
RTP/NIEHS
New Hanover County
New Hanover County
Wrightsville Beach*
Auburn
Candia
Canterbury*
Claremont
City
Alexandria
Duluth
Elk River
Fergus Falls
Fosston
Mankato
Minneapolis
Perham
Red Wing
Red Wing
Rochester
Scott
Thief River Falls
Thief River Falls
Ft. Leonard Wood
Moss Point
Livingston
Charlotte
Cherry Point
RTP
Wilmington
Wilmington
Wrightsville Beach
Auburn
Candia
Canterbury
Claremont
State
MN
MN
MN
MN
MN
MN
MN
MN
MN
MN
MN
MN
MN
MN
MO
MS
MT
NC
NC
NC
NC
NC
NC
NH
NH
NH
NH
Capacity
tons/yr
26,280
146,000
547,500
34,310
29,200
262,800
438,000 .
41,610
262,800
26,280
73,000
21,900
1,825
36,500
28,470
54,750
26,280
85,775
365
14,600
73,000
91,250
18,250
1,825
5,475
3,650
73,000
Mg/yr
23,891
132,727
497,727
31,191
26,545
238,909
398,182
37,827
238,909
23,891
66,364
19,909
1,659
33,182
25,882
49,773
23,891
77,977
332
13,273
66,364
82,955
16,591
1,659
4,977
3,318
66,364
a No longer operational.
June 1996
A-15
SAB REVIEW DRAFT
-------
Table A-9 (continued)
Existing MWC Facilities (as of December, 1991)
Facility
Concord
Durham
Lincoln
Litchfield
Nottingham
Pelham "
Pittsfield*
Plymouth
Waton
Wolfeboto
Atlantic County Jail
Camden County
FortDix
Essex County
Warren County
Union County
Gloucester County
Albany*
Babylon
SW Brooklyn*
Cattaraugus*
Ellis Island
Fire Island
Glen Cove*
Hempstead
Huntington
Islip
City
Concord
Durham
Lincoln
Litchfield
Nottingham
Pelham
Pittsfield
Plymouth
Waton
Wolfeboro
Atlantic County Jail
Camden
FortDix
Newark
Oxford Township
Rab.way
West Deptford
Albany
Babylon
Brooklyn
Cuba
Ellis Island
Fire Island
Glen Cove
Hempstead
Huntington
Islip
State
NH
NH
NH
NH
NH
NH
NH
NH
NH
NH
NJ
NJ
NJ
NJ
NJ
NJ
NJ
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
Capacity
tons/yr
182,500
39,420
8,760
8,030
2,920
8,760
17,520
5,840
10,950
5,840
5,110
383,250
29,200
831,105
146,000
525,600
209,875
219,000
273,750
350,400
40,880
4,380
4,380
91,250
846,435
273,750
189,070
Mg/yr
165,909
35,836
7,964
7,300
2,655
7,964
15,927
5,309
9,955
5,309
4,645
348,409
26,545
755,550
132,727
477,818
190,795
199,091
248,864
318,545
37,164
3,982
3,982
82,955
769,486
248,862
171,884
a No longer operational.
June 1996
A-16
SAB REVIEW DRAFT
-------
Table A-9 (continued)
Existing MWC Facilities (as of December, 1991)
Facility
Liberty Island
Long Beach
Belts Ave (NY City)a
Niagara Falls-Occidental
Oswego County
Westchester County
Dutchess County
Oneida County
Akron
Columbus*
South Montgomery County
North Montgomery County
Euclid*
Miami
Poteau
Tulsa
Wilburtona
Bendon
Marion County
Coos Bay
Courthouse-Coquille
Courthouse-Coquille
Delaware County (Chester)
Lancaster County
Westmoreland County
Harrisburg
York County
City
Liberty Island
Long Beach '
New York
Niagara Falls
Oswego County
PeekskiU
Poughkeepsie
Rome
Akron
Columbus
Dayton
Dayton
Euclid
Miami
Poteau
Tulsa
Wilburton
Benton
Brooks
Coos Bay
Coquille
Coquille
Chester
Conoy Township
Greensburg
Harrisburg
Manchester Township
State
NY
NY
NY
NY
NY
NY
NY
NY
OH
OH
OH
OH
OH
OK
OK
OK
OK
OR
OR .
OR
OR
OR
PA
PA
PA
PA
PA
Capacity
tons/yr
4,380
• 73,000
365,000
730,000
73,000
821,250
146,000
73,000
365,000
730,000
328,500
328,500
73,000
38,325
9,125
273,750
6,570
9,490
200,750
36,500
54,750
4,745
981,120
438,000
18,250
262,800
490,560
Mg/yr
3,982
66,364
331,818
663,636
66,364
746,591
132,727
66,364
331,818
663,636
298,636
298,636
66,364
34,841
8,295
248,864
5,973
8,627
182,500
33,182
49,773
4,314
891,927
398,182
16,591
238,909
445,964
a No longer operational.
June 1996
A-17
SAB REVIEW DRAFT
-------
Table A-9 (continued)
Existing MWC Facilities (as of December, 1991)
Facility
Philadelphia ECa
Philadelphia NW*
Montgomery County
Charleston County
Hampton
Davidson County
Dyersburg*
Galletin
Lewisburg*
Nashville
Carthage City
Center
Clebume
Gatesville (Prison)*
Walker County*
Walker County (Prison)*
Grimes County*
Anderson County
Quitman
Waxahachie*
Davis County
Alexandria
Arlington (Pentagon)
Galaxa
Hampton
Harrisonburg
Fairfax County
City
Philadelphia
Philadelphia
Plymouth Township
Charleston
Hampton
Davidson County
Dyersburg
Galletin
Lewisburg
Nashville
Carthage
Center
Clebume
Gatesville
Huntsville
Huntsville
Navasota
Palestine
Quitman
Waxahachie
Layton
Alexandria
Arlington
Galax
Hampton
Harrisonburg
Lorton
State
PA
PA
PA
SC
SC
TN
TN
TN
TN
TN
TX
TX
TX
TX
TX
TX
TX
TX
TX
TX
UT
VA
VA
VA
VA
VA
VA
Capacity
tons/yr
273,750
273,750
438,000
219,000
87,600
67,525
36,500
73,000
21,900
408,800
14,600
14,600
41,610
7,300
9,125
9,125
9,125
9,125
7,300
18,250
146,000
355,875
18,250
20,440
73,000
36,500
1,095,000
Mg/yr
248,864
248,864
398,182
199,091
79,636
61,386
33,182
66,364
19,909
371,636
13,273
13,273
37,827
6,636
8,295
8,295
8,295
8,295
6,636
16,591
132,727
323,523
16,591
18,582
66,364
33,182
995,455
a No longer operational.
June 1996
A-18
SAB REVIEW DRAFT
-------
Table A-9 (continued)
Existing MWC Facilities (as of December, 1991)
Facility
Norfolk Navy Yard
Norfolk Naval Station1
Salem*'
Readsboro
Rutland4
Stamford
Bellingham
Fort Lewis
Friday Harbor
Skagit County
Spokane
Tacoma
Barron Co.
La Crosse County*
Madison (Power Plant)3
Muscodaa
Waukesha*
Total
City
Norfolk
Norfolk
Salem
Readsboro
Rutland
Stamford
Ferndale
Fort Lewis
Friday Harbor
Mt. Vernon
Spokane
Tacoma
Almena
La Crosse
Madison
Muscoda
Waukesha
State
VA
VA
VA
VT
VT
VT
WA
WA
WA
WA
WA
WA
WI
WI
WI
WI
WI
Capacity
tons/yr
730,000
131,400
36,500
4,745
87,600
3,650
36,500
43,800
4,380
64,970
292,000
109,500
29,200
146,000
43,800
43,800
64,240
41,729,720
Mg/yr
663,636
119,455
33,182
4,314
79,636
3,318
33,182
39,818
3,982
59,064
265,455
99,545
26,545
132,727
39,818
39,818
58,400
37,936,109
Source: Memorandum from Fenn, D., and K. Nebel, Radian Corporation, to Stevenson, W., U. S. Environmental
Protection Agency. March 9, 1992.
a No longer operational.
June 1996
A-19
SAB REVIEW DRAFT
-------
Table A-10
Mercury Emissions From MWCs by Combustor Type For 1994
Combustor
type
Mass burn
Mass burn
Mass bum
Mass burn
Mass bum
RDF
Modular
Modular
Unknown
Total
Control status
U
SD
DSI
ESP
a
SD
ESP
SD
U
Process rate,1
106Mg/yr
0.43
10.8
0.85
63
1.5
8.1
1.0
0.73
0.03
Uncontrolled
emission
factor, g/Mg
2.82
2.82
2.82
2.82
2.82
2.77
2.82
2.82
2.82
Control
efficiency, %
0
50
50
0
85
50
0
50
0
Annual Emissions
Mg/yr
1.2
15.2
1.2
17.8
0.6
11.2
2.8
1.0
0.07
51.1
Tons/yr
1.3
16.7
1.3
19.6
0.7
12.3
3.1
1.1
0.08
56.2
Key:
SD = Spray dryer with either ESP or fabric filter
ESP - Electrostatic precipitator
DSI = Duct sorbent injection with either ESP or fabric filter
U = Uncontrolled
CI = Activated carbon injection
a Annual totals assume that plants operate at 85% of design capacity.
Basis of Input Data for EPA's Emissions Calculations
1. Under the assumption that ESP's provide essentially no control, the facility-average concentrations at 7 percent oxygen
for uncontrolled and ESP-controlled mass burn (including modular) and RDF systems were averaged to obtain the
following "typical" concentrations:
Mass Burn - 696 ug/dscm
RDF - 561
2. The F-factor for municipal waste combustors was assumed to be 0.257 x 10'6 dscm/J at 0 percent oxygen and the
heating values were assumed to be 10,500 kJ/kg for MSW and 12,800 U/kg for RDF. The F-factor was converted from
0 percent oxygen to 7 percent oxygen (at which concentrations are based) using a factor of 1.5.
3. Spray dryer or dry sorbent injection systems combined with fabric filters or ESP's and wet scrubber systems are assumed
to achieve 50 percent removal. No other control measures achieve appreciable mercury control.
Calculations
Uncontrolled Emission Factors
• Mass bum/modular - 696 ug/dscm * 0.257 x Iff6 dscm/J * 10,500 kJ/kg * 1.5 = 2.82 g/Mg
• RDF - 561 ug/dscm * 0.257 x 10'6 dscm/J * 12,800 kJ/kg * 1.5 = 2.77 g/Mg
Controlled Emissions
Annual Emissions
= Process rate * emission factor * (100-efficiency)
100
Source: Locating and Estimating Air Emissions from Sources of Mercury and Mercury Compounds. U.S. Environmental
Protection Agency. July 1993.
June 1996
A-20
SAB REVIEW DRAFT
-------
Industry Estimates for MWC Mercury Emissions
Industry estimates of mercury emissions for the MWC source category are 40 Mg/year (44 tons/year) (Kiser, 1991). The
differences between EPA's emission estimates and the industry estimates stem from essentially three differences in the data
assumed for the emissions calculations. The basis for the industry estimate is as follows.
Basis of Input Data for Industry's Emissions Calculations
1. The average level of mercury from MWCs was assumed to be 500 ug/dscm compared to 696 ug/dscm that EPA used for
mass-burn combustors and 561 ug/dscm that EPA used for RDF combustors.
2. Industry estimates attribute a 10 percent reduction in mercury for MWCs equipped with an ESP. EPA estimates do not
credit any mercury removal to MWCs equipped with ESPs.
3. Industry estimates attribute an 80 percent reduction in mercury for MWCs equipped with a spray dryer and firing RDF.
EPA credits a 50 percent reduction for these facilities.
June 1996 A-21 SAB REVIEW DRAFT
-------
Table A-ll
MWI Population By State
State
Alabama
Alaska
Arizona
Arkansas
California
Colorado
Connecticut
Delaware
Florida
Georgia
Hawaii
Idaho
Illinois
Indiana
Iowa
Kansas
Kentucky
Louisiana
Maine
Maryland
Massachusetts
Michigan
Minnesota
Mississippi
Missouri
Montana
Nebraska
Nevada
New Hampshire
New Jersey
New Mexico
New York
North Carolina
North Dakota
Ohio
Oklahoma
Oregon
Pennsylvania
Rhode Island
South Carolina
Commercial units
No.
2
1
1
1
10
2
la
0
12
- 1
0
ND
2
4
0
Some
1
1
1
ND
3
1
0
0
2
ND
0
0
1
1
3
2
7
0
15
1
2
8
1
3
Capacity range, Ib/hr
Up to 60,000 Ib/wk
ND
ND
ND
ND
225
1,200
NA
300-3,750
ND
NA
ND
ND
ND
NA
ND
1,500
1,500
150
ND
350-UOO
720
NA
NA
ND
ND
NA
NA
ND
5,000
75-1,000
1,950
150-3,250
NA
500-1,700
ND
200-1,000
ND
975
6,250
Onsite units
No.
250
10 to 12
97
150
157
46
44b
20
273
ND
6or7
_20-25
permitted
259
91
ND
ND
ND
100-125
22
121
200
160
145
125-175
100
<50
80
17
27
154
31
599
29
50
125
93
31
186
11
70
Capacity range,
Ib/hr
ND
ND
ND
ND
ND
13-1,000
20-1,500
ND
ND
* ND
100-1,000
ND
2-1,500
Most <7 tons/d
<125
ND
ND
<500 to >1,000
(most <500)
20-1,000
ND
3-1,875
ND
50-1,250
ND
Most <500
ND
ND
40-360
Most 75-150
20-1,560
25-360
3-3,000
60-2,100
Most <500
25-2.500
ND
25-750
ND
50-1..500
<500 to 1,000
Facilities
included
H,N
H
All
ND
H,N,O
H
All
H,V
H,F,V,A4-,0
H
H
H,F,V,A^,0
H
H
H
H,N,L
All
All
H,N
H
All
H
ND
H
H
H
All
H,F,V,AX
All
H,L
H
H,F,V,AJ-,O
UNO.
H
H
H
H
a Closed as of 1994.
b Twenty-two facilities as of 1994.
June 1996
A-22
SAB REVIEW DRAFT
-------
Table A-ll (continued)
MWI Population By State
State
South Dakota
Tennessee
Texas
Utah
Vermont
Virginia
Washington
West Virginia
Wisconsin
Wyoming
Commercial units
No.
0
>1
2
0
0
Ior2
6
ND
4
0
Capacity range, Ib/hr
NA
ND
ND
NA
ND
ND
1,600-7,500
ND
\ND
-NA
Onsite units
No.
30
126
ND
20 .
9
ND
137
50
ND
30
Capacity range,
Ib/hr
ND
ND
ND
<500
ND
ND
40-825
ND
ND
ND
Facilities
included
H
H
H
H
All
H
. H
KEY:
General
ND = no data
NA = not applicable
Facility types
H = hospital/medical centers
F = funeral homes
V a veterinaries
A = animal shelters
L = laboratories, clinical and research
N = nursing homes
O = other/unidentified facilities
All = all MWI facilities (using current broad working definition)
Source: U.S. EPA, 1994. Medical Waste Incinerators-Background Information for Proposed Standards and
Guidelines: Industry Profile Report for New and Existing Facilities. EPA-453/R-94-042a. Research
Triangle Park, NC.
June 1996
A-23
SAB REVIEW DRAFT
-------
Table A-12
Mercury Emissions From Model Medical Waste Incinerators
Model
Number
1
2
3
4
5
6
7
Type of Unit
Continuous
Continuous
Intermittent
Intermittent
Intermittent
Batch v
Pathological
Operating
Hours/year
7,760
3,564
4,212
4,212
3,588
3,520
2,964-
Total
Number
of Units
154
182
171.
742
2,097
335
1,305
4,986
Waste feed capacity
Mg/yr
0.68
0.45
0.68
0.27
0.09
0.23
0.09
2.5
Tons/yr
0.75
0.5
0.75
0.3
0.1
0.25 .
0.1
2.75
Annual Emissions
Mg/yr
14.93
5.40
9.00
15.62
12.54
1.05
0.12
58.7
Tons/yr
16.47
5.96
9.93"
17.23
13.83
1.16
0.13
64.71
June 1996
A-24
SAB REVIEW DRAFT
-------
Table A-13
U.S. Sewage Sludge Incinerators
Facility
Anchorage
Petersburg
Wrangell
Bars tow
Lake Arrowhead
Martinez
Palo Alto
Redwood City
Sacramento
San Mateo
South Lake Tahoe
Tahoe Trackee
Central Contra Costa
Yosemite
Mattabassett
Mattabassett
Hartford WPCF
New Canaana
East Shore WPCF
New London WPCF
Norwalka
Stamford
Waterbury WPCF*
West Haven
Willimantic WPCF
Jacksonville
Pensacola WWTP
R.M. Clayton WWTP
Location
Anchorage
Petersburg
Wrangell
Bars tow
Lake Arrowhead
Martinez
Palo Alto
Redwood City
Sacramento
San Mateo
South Lake Tahoe
Truckee
Walnut Creek
Yosemite National Park
Cromwell
Cromwell
Hartford
New Canaan
New Haven
New London
Norwalk
Stamford
Waterbury
West Haven
Willimantic
Jacksonville
Pensacola
Atlanta
State
AK
AK
AK
CA
CA
CA
. CA
CA
CA
CA
CA
CA
CA
CA
CT
CT
CT
CT
CT
CT
CT
CT
CT
CT
CT
FL
FL
GA
Ca]
Ton/yr
401.5
N/A
73
912.5
1,752
N/A
2,190
N/A
2,628
1,788.5
292
N/A
16,060
1,241
4,562.5
N/A
122,640
14,198.5
10,658
18,907
13,140
34,565.5
2,372.5
N/A
N/A
10,366
N/A
N/A
>acity
•Mg/yr
365
N/A
66.4
829.5
1,592.7
N/A
1,990.9
N/A
2,389.1
1,625.9
265.5
N/A
14,600
1,128.2
4,147.7
N/A
111,490.9
12,907.7
9,689.1
17,188.2
11,945.5
31,423.2
2,156.8
N/A
N/A
9,423.6
N/A
N/A
a Not operational as of 1994.
June 1996
A-25
SAB REVIEW DRAFT
-------
Table A-13 (continued)
U.S. Sewage Sludge Incinerators
Facility
Atlanta (Utoy)
Atlanta (Bolton Rd)
Decatur
Gainesville
Cobb County
Savannah
San Island WWTF
Honouliuli WWTP
Oahu
Cedar Rapids WPCF
Davenport
Dubuque
DecatnrSTP
Indianapolis-Belmont
Kansas City
Kaw Point
Mission Township STP
Turkey Creek MSD #1
Cynthiana
Kenton County
New Orleans West Bank
STP
Lake Charles
Lake Charles-Plant B
Lake Charles-Plant C
East Bank STP #2
Natchitoches
East Bank STP #1
Attleboro Advanced WTF
Chicopee
Location
Atlanta
Atlanta
Decatur
Gainesville
Marietta
Savannah
Honolulu
Honouliuli
Oahu
Cedar Rapids
Davenport
Dubuque
Decatur
Indianapolis
Kansas City
Kansas City
Mission
Shawnee Mission
Cynthiana
Kenton
Algiers
Lake Charles
Lake Charles
Lake Charles
Lake Charles
Natchitoches
New Orleans
Attleboro
Chicopee
State
GA
GA
GA
GA
GA
GA
ffl
ffl
HI
IA
IA
IA
IL
IN
KS
KS
KS
KS
KY
KY
LA
LA
LA
LA
LA
LA
LA
MA
MA
Capacity
Ton/yr
2,956.5
47,304
16,352
2,007.5
7,227
4,380
9,453.5
N/A
N/A
8,869.5
12,994
20,440
N/A
132,458.5
6,570
14,600
N/A
6,497
N/A
N/A
N/A
2,190
N/A
N/A
14,965
N/A
10,950
N/A
2,628
Mg/yr
2,687.7
43,003.6
14,865.5
1,825
6,570
3,981.8
8,594.1
N/A
N/A
8,063.2
11,812.7
18,581.8
N/A
120,416.8
5,972.7
13,272.7
N/A
5,906.4
N/A
N/A
N/A
1,990.9
N/A
N/A
13,604.5
N/A
9,954.5
N/A
2,389.1
June 1996
A-26
SAB REVIEW DRAFT
-------
Table A-13 (continued)
U.S. Sewage Sludge Incinerators
Facility
Chicopee
Fall River
Htchburg East WWTP
Lynn
Upper Blackstone WWTP
New Bedford WWTP
Greater Lawrence SD
WWTP
Annapolis City SIP
Patapsco
Ocean City
Cox Creek WWTP
Ann Arbor
Battle Creek
Bay City STP
Bay County STP
Detroit (1)
Detroit (2)
East Lansing
Grand Rapids
Kalamazoo WWTP
Lansing WWTP
NilesWWTP
Owosso WWTP
Pontiac STP
Port Huron
Trenton WWTP
Warren
Wyandotte STP
Location
Chicopee
Fall River
Fitchburg
Lynn
Millbury
New Bedford
North Andover
Annapolis
Baltimore
Ocean City
Riviera Beach
Ann Arbor
Battle Creek
Bay City
Bay County
Detroit
Detroit
East Lansing
Grand Rapids
Kalamazoo
Lansing
Niles
Owosso
Pontiac
Port Huron
Trenton
Warren
Wyandotte
State
MA
MA
MA
MA
MA
MA
MA
MD
MD
MD
MD
MI
MI
MI
MI
MI
MI
MI
MI
MI
MI
MI
MI
MI
MI
MI
MI
MI
Caj
Ton/yr
2,628
N/A
14,198.5
N/A
12,811.5
5,913
33,142
N/A
35,916
2,920
N/A
19,710
N/A
1,168
N/A
148,920
245,937
11,826
-11,826
17,520
N/A
N/A
N/A
23,652
2,774
N/A
9,453.5
88,768
jacity
Mg/yr
2,389.1
N/A
12,907.7
N/A
11,646.8
5,375.5
30,129.1
N/A
32,650.9
2,654.5
N/A
17,918.2
N/A
1,061.8
N/A
135,381.8
223,579.1
10,750.9
10,750.9
15,927.3
N/A
N/A
N/A
21,501.8
2,521.8
N/A
8,594.1
80,698.2
June 1996
A-27
SAB REVIEW DRAFT
-------
Table A-13 (continued)
U.S. Sewage Sludge Incinerators
Facility
Ypsilanti Community
WWTP
Duluth
Metropolitan TP
Seneca TP
Independence
Kansas City
Little Blue Valley
St. Louis (Lenay STP)
St Louis (Bissel Point
STP)
Greensboro
Manchester WWTP
Rocky Mount
Shelby
Lebanon WWTP
Merrimack WWTP
Atlantic City
Gloucester Township
Somerset Raritan Valley
Authority
West Side STP
Two Bridges
Parsippany
Rockaway Valley
Stony Brook RSA STP
#1
Bayshore Regional Sewer
Authority
NW Bergen County
Utilities
Wayne
Location
Ypsilanti
Duluth
St. Paul
St. Paul
Independence
Kansas City
Little Blue
St. Louis
St Louis
Greensboro
Manchester
Rocky Mount
Shelby
Lebanon
Merrimack
Atlantic City
Blackwood
Bridgewater
Jersey City
Lincoln Park
Parsippany
Parsippany-Troy Hills
Princeton
Union Beach
Waldwick
Wayne
State
MI
MN
MN
MN
MO
MO
MO
MO
MO
NC
NC
NC
NC
NH
NH
NJ
NJ
NJ
NJ
NJ
NJ
NJ
NJ
NJ
NJ
NJ
Cai
Ton/yr
19,710
12,410
283,824
7,081
3,540.5
16,371
N/A
53,217
118,260
16,571
N/A
2,737.5
5,913
2,628
N/A
9,453.5
3,504
5,110
5,037
24,i[)90
28,397
N/A
14,417.5
10950
6,570
35,040
>acity
Mg/yr
17,918.2
11,281.8
258,021.8
6,437.3
3,218.6
15,064.5
N/A
48,379.1
107,509.1
15,064.5
N/A
2,488.6
5,375.5
2,389.1
N/A
8,594.1
3,185.5
4,645.5
4,579.1
21,900
25,815.5
N/A
13,106.8
9,954.5
5,972.7
31,854.5
June 1996
A-28
SAB REVIEW DRAFT
-------
Table A-13 (continued)
U.S. Sewage Sludge Incinerators
Facility
Mountain View Sewer
Authority
Round Hill
Douglas County SID #1
WWTF
Albany (North)
AlbanyVSouth)
Amherst
Arlington
Auburn
Bath
Beacon WPCP
Birds Island STP
Southtowns Advanced
WWTF
Dunkirk STP
Numburg
Glen Cove
Glens Falls
NW Quadrant TP
Little Falls
Mamaroneck
New Rochelle SD STP
Niagra County
Utica
Orangetown DPW
East STP
West STP
Port Chester SDSTP
Port Washington
Gates Chile Ogden STP
Location
Wayne Township
Round Hill Village
Zephyr Cove-Round Hill
Village
Albany
Albany
Amherst
Arlington
Auburn
Bath
Beacon
Buffalo
Buffalo
Dunkirk
Erie County
Glen Cove
Glens Falls
Greece
Little Falls
Mamaroneck
New Rochelle
Niagara County
Oneida County
Orangetown
Oswego
Oswego
Port Chester
Port Washington
Rochester
State
NJ
NV
NV
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
NY
Capacity
Ton/yr
N/A
3,285 .
1,825
47,304
33,507
N/A
3,066
14,782.5
N/A
3,540.5
67,014
15,768
N/A
105,120
9,125
N/A
N/A
1,423.5
N/A
N/A
N/A
21,900
6,132
N/A
N/A
N/A
N/A
13,140
Mg/yr
N/A
2,986.4
1,659.1
43,003.6
30,460.9
N/A
2,787.3
13,438.6
N/A
3,218.6
60,921.8
14,334.5
N/A
95,563.6
8,295.5
N/A
N/A
1,294.1
N/A
N/A
N/A
19,909.1
5,574.5
N/A
N/A
N/A
N/A
11,945.5
June 1996
A-29
SAB REVIEW DRAFT
-------
Table A-13 (continued)
U.S. Sewage Sludge Incinerators
Facility
Rochester (NW Quad)
Frank E. Van Lare
WWTP
Saratoga
Schenectady STP
Two Mile Creek STP
Watertown
Watertown
Akron WWTP
Canton WWTP
Little Miami WWTP
Cincinnati (Millcreek)
Cleveland (Southerly
WWTP)
Cleveland (Westerly
STP)
Columbus (South)
Columbus (Jackson Pike
WWTP)
Euclid WWTP
Warren County
Lorain
Willoughby-Eastlake
WWTP
Youngstown WWTP
Tigard
Ambridge STP
Kiski Valley WPCA
Bridgeport SIP
Delcora-Chester STP
Hatfield Township STP
Location
Rochester
Rochester
Saratoga
Schenectady
Tonawanda
Watertown
Watertown
Akron
Canton
Cincinnati
Cincinnati
Cleveland
Cleveland
Columbus
Columbus
Euclid
Franklin
Lorain
Willoughby
Youngstown
Tigard
Ambridge
Appollo
Bridgeport
Chester
Colmer
State
NY
NY
NY
NY
NY
NY
NY
OH
OH
OH
OH
OH
OH
OH
OH
OH
OH
OH
OH
OH
OR
PA
PA
PA
PA
PA
Capacity
Ton/yr
17,520
26,280
N/A
51,100
N/A
N/A
' 7,665
14,162
18,250
N/A
61,466
94,608
70,956
16,571
14,198.5
•
7,884
N/A
N/A
7,665
14,782.5
5,475
N/A
49,676.5
N/A
7,081
2,080.5
Mg/yr
15,927.3
23,890.9
N/A
46,454.5
N/A
N/A.
6,968.2
12,874.5
16,590.9
N/A
55,878.2
86,007.3
64,505.5
15,064.5
12,907.7
7,167.3
N/A
N/A
6,968.2
13,438.6
4,977.3
N/A
45,160.5
N/A
6,437.3
1,891.4
June 1996
A-30
SAB REVIEW DRAFT
-------
Table A-13 (continued)
U.S. Sewage Sludge.Incinerators
Facility
Duryea
Erie
Hershey
• City of Johnstown
Cumberland City
E. Morristown Plymouth
TP
Upper Gwynedd
Lower Lackawanna STP
Alcosan WWTP
Tyrone
Trout Run WPCC
Hazeltown
Wyoming Valley
Sanitation Authority
Upper Moreland-Hatboro
TP
York
Cranston
Hanisburg
Providence
Charleston
Columbia
North Charleston
Bristol
Maryville Regional STP
Central WWTP
Newport
Alexandria STP
Arlington COWPCP
Fairfax
Location
Duryea
Erie
Hershey
Johnstown
Lemoyne
Norristown
North Wales
Old Forge
Pittsburgh
Tyrone
Upper Merion Township
West Hazeltown
Wilkes-Barre
Willow Grove
York
Cranston
Harrisburg
Providence
Charleston
Columbia
North Charleston
Bristol
Maryville
Nashville
Newport
Alexandria
Arlington
Fairfax
State
PA
PA
PA
PA
PA
PA
PA
PA
PA
PA
PA
PA
PA
PA
PA
RI
RI
RI
SC
SC
SC
TN
TN
TN
TN
VA
VA
VA
Caj
Ton/yr
9,453.5
49,275
14,782.5
2,956.5
N/A
N/A
401.5
N/A
N/A
1,861.5
N/A
1,642.5
11,826
N/A
14,162
7,446
49,421
N/A
11,826
2,993
N/A
5,913
4,745
33,580
2,847
N/A
12,702
23,652
jacity
Mg/yr
8,594.1
44,795.5
13,438.6
2,687.7
N/A
* N/A
365
N/A
N/A
1,692.3
N/A
1,493.2
10,750.9
N/A
12,874.5
6,769.1
44,928.2
N/A
10,750.9
2,720.9
N/A
5,375.5
4,313.6
30,527.3
2,588.2
N/A
11,547.3
21,501.8
June 1996
A-31
SAB REVIEW DRAFT
-------
Table A-13 (continued)
U.S. Sewage Sludge Incinerators
Facility
Fairfax (Lower Potomac
STP)
Hopewell
Boat Harbor
Lamberta Point WPCF
Army Base WWTP
(Hampton Rds.)
Chesapeake-Elizabeth
WPCF
Williamsburg WPCF
Potomac River STP
Edmonds
Lynnwood
Vancouver
Brookfield STP
Green Bay WWTP
Milwaukee
Clarksburg
Huntington
Total
Location
Fairfax
Hopewell
Newport News
Norfolk
Norfolk
Virginia Beach
Williamsburg
Woodbridge
Edmonds
Lynnwood
Vancouver
Brookfield
Green Bay
Milwaukee
Clarksburg
Huntington
State
VA
VA
VA
VA
VA
VA
VA
VA
WA
WA
WA
WI
WI
WI
WV
WV
Capacity
Ton/yr
33,142
2,956.5
12,300.5
21,133.5
9,307.5
t>
8,322
20,330.5
N/A
584
255.5
12,410
1,423.5
31,937.5
2,591.5
N/A
N/A
3,208,240.5
Mg/yr
. 30,129.1
2,687.7
11,182.3
19,212.3
8,461.4
7,565.5
18,482.3
N/A
530.9
232.3
11,281.8
1,294.1
29,034.1
2,355.9
N/A
N/A
2,916,582.3
Source: Locating and Estimating Air Toxic Emissions from Sewage Sludge Incinerators. U.S. EPA.
EPA-450/2-90-009. May 1990.
June 1996
A-32
SAB REVIEW DRAFT
-------
APPENDIX B
MERCURY REMOVAL CAPABILITIES OF
PARTICULATE MATTER AND ACID GAS CONTROLS FOR UTILITIES
-------
-------
APPENDIX B
MERCURY REMOVAL CAPABILITIES OF PARTICIPATE
MATTER AND ACID GAS CONTROLS FOR UTILITIES
Existing air pollution control devices (APCDs) on utilities typically control either paniculate
matter (PM) or sulfur dioxide (SO2) emissions, or both. Nitrogen oxides may be controlled by an
APCD, but are usually controlled by combustion modification. Generally, a wet scrubber is used to
control SO2 emissions only, while a dry scrubber can control SO2 emissions and PM because it is
usually built with a 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 SO2 emissions). Most utility boilers are equipped with an ESP or FF
before the wet FGD units to collect PM.
Figure B-l shows the relationship between mercury removal and the inlet temperature for wet
FGD devices. Table B-l summarizes available test data for FGD units. FGDs have a median mercury
removal efficiency of about 17.3 percent, with a range from 0 percent to 59.3 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.
June 1996 B-l SAB REVIEW DRAFT
-------
Figure B-l
Removal of Mercury By An FGD (Coal)
100 -i
80
60 -
0>
CL
I
J£
S.
40 —
20 -
260
280
300
320
340
FGD Inlet Temperature (F)
June 1996
B-2
SAB REVIEW DRAFT
-------
Table B-l
Test Data for FGD Units
Unit
EPRI Site 11
EPRI Site 12
NSP Sherbume 1 & 2 Test A
NSP Sherburne 1 & 2 Test B
DOE Yates
DOE Coal Creek
Control Device
Wet limestone FGD (inlet Hg
concentration of 9.9 ug/dscm)
Wet limestone FGD
Wet limestone FGD (inlet Hg
concentration of 8.1 ug/dscm)
Wet limestone FGD (inlet Hg
concentration of 11.6 ug/dscm)
Wet limestone and jet bubbling
reactor FGD (inlet Hg concentration
of 6.0 ug/dscm)
Wet lime FGD (inlet Hg
concentration of 10.0 ug/dscm)
Median
Mean
Standard deviation
Hg Removal %
10.87
0*
22.63
59.3
- 45.91
12.05
1734
25.13
22.85
Reference
Radian, 1993a
Radian, 1993b
Interpoll,
1990a
Interpoll, 1991
EPRI, 1993a
Battelle, 1993a
a This unit was re-tested for mercury as part of a ESP/FGD system. Since there was no way of determining
which 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 paniculate
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 hi 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.
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3
o
I
Figure B-2
Removal of Mercury By A Spray Dryer Adsorber/
Fabric Filter (Coal)
100 -i
80 —
60 -
40 —
20 -
270
280 290 300
SDA Inlet Temperature (F)
310
Table B-2
Spray Dryer Adsorption Data
Unit
EPRI Site 14
DOE Springerville
Sherbume 3 Test A
Sherbume 3 Test B
*
Control Device
SDA/FF (inlet Hg concentration of 1.0
|ag/dscm)
SDA/FF (inlet Hg concentration of 8.3
Hg/dson)
SDA/FF (inlet Hg concentration of 6.8
Hg/dson)
SDA/FF (inlet Hg concentration of 13.4
Hg/dsan)
Median
Mean
Standard deviation
Hg Removal %
0
2.16
45.71
54.5
23.94
25.59
28.54
Reference
Radian, 1993c
Southern Research
Institute, 1993a
Interpoll, 1990b
Interpoll, 1991
June 1996
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B J 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.
Figure B-3
Removal of Mercury By A FF (Coal)
100 -,
80 —
60 -
40
20 -
96
» I p
97 98 99
PM Collection Efficiency (Percent)
100
June 1996
B-5
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Table B-3
Fabric Filter Data
Unit
EPRI Site 13
EPRI Site 115
NSP Riverside 6 & 7
DOE Niks #2 w/NOx
DOE Boswell
Control Device
FF (inlet Hg concentration of
0.3 ug/dscm)
FF (inlet Hg concentration of
1.8 ug/dscm)
FF (inlet Hg concentration of
4.8 ug/dscm)
FF (inlet Hg concentration of
25.8 ug/dscm)
FF (inlet Hg concentration of
6.4 ug/dscm)
Median
Mean
Standard deviation
Hg Removal %
0
73.36
0
8.39
60.59
8.39
28.47
35.61
Reference
Radian, 1993d
Camot, 1994a
InterpoU, 1992a
Battelle, 1993b
Weston, 1993a
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 hot-side 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 hi 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.
June 1996
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Figure B-4
Removal of Mercury By Electrostatic Precipitators (Cold-Side, Coal)
100 -i
80 —
5
o
Q>
o:
8
£
60 -
40 —
20 -
92 94 96 98
PM Collection Efficiency (Percent)
100
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Table B-4
Test Data for Cold-Side Electrostatic Precipitators (Controlling Coal-Fired Units)
Unit
EPRI Site 11
EPRI Site 12
EPRI Site 15
EPRI Site 102
NSP High Bridge 3,4,5,6
NSP High Bridge 1,3,4
NSP Black Dog #2
NSP Riverside #8
EPRI Site 114 /Test A
EPRI Site 1 14 / Test B
DOE Niles #2
DOEYates
DOE Coal Creek
EPRI Site 16/OFA/LNOx Burners
EPRI Site 16/OFA
DOE Cardinal
DOE Baldwin
Control Device
ESP, CS (inlet Hg concentration of 3.4 pg/dscm)
ESP, CS (inlet Hg concentration of 9.1 ng/dscm)
ESP, CS (inlet Hg concentration of 4.9 jug/dscm)
ESP, CS (inlet Hg concentration of 9.0 ng/dscm)
ESP, CS (inlet Hg concentration of 4.4 jjg/dscm)
ESP, CS (inlet Hg concentration of 5.1 ng/dscm)
ESP, CS (inlet Hg concentration of 2.8 jjg/dscm)
ESP, CS (inlet Hg concentration of 2.9 ng/dscm)
ESP, CS (inlet Hg concentration of 10.6 jig/dscm)
ESP, CS (inlet Hg concentration of 10.6 jig/dscm)
ESP, CS (inlet Hg concentration of 24.7 jig/dscm)
ESP, CS (inlet Hg concentration of 5.9 ng/dscm)
ESP, CS (inlet Hg concentration of 11.0 ng/dscm)
ESP, CS (inlet Hg concentration of 11.5 ng/dscm)
ESP, CS (inlet Hg concentration of 7.6 jig/dscm)
ESP, CS (inlet Hg concentration of 2.3 ng/dscm)
ESP, CS (inlet Hg concentration of 7.0 ng/dscm)
Median
Mean
Standard deviation
Hg
Removal
%
0
82.35
0
0
6.87
8.21
21.56
0
29.8
16.16
26.55
55.23
13.15
54.8
9.38
73.9
26.13
16.16
24.95
2633
Reference
Radian, 1993a
Radian, 1993b
Radian, 1992a
Radian, 1993e
Interpoll,
1992b
Interpoll,
1992c
Interpoll,
1992d
Interpoll,
1992e ||
Radian, 1994a II
Radian, 1994a 1
Battelle, 1993
-------
Figure B-5
Removal of Mercury By Electrostatic Precipitators (Hot-Side, Coal)
100 -i
80 —
s
o
I
60 -
40
OJ
Q_
20 -
98
99
100
PM Collection Efficiency (Percent)
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Table B-5
Test Data for Hot-Side Electrostatic Precipitators (Controlling Coal-Fired Units)
Unit
EPRI Site 110
EPRI Site 110 with
NOX control
Control Device
ESP, HS (inlet Hg concentration of
5.3 ug/dscm)
ESP, HS (inlet Hg concentration of
0.3 ng/dson)
. Median
Hg Removal %
0
0
0 (see description
in text)
Reference
Southern Research
Institute, 1993b
Southern Research
Institute, 1993b
x Figure B-6
Removal of Mercury By Electrostatic Precipitators (Oil)
100 -i
80 —
60 -
0)
>_
§ 40 -
-------
Table B-6
Test Data for Cold-Side Electrostatic Precipitators (Controlling Oil-Fired Units)
Unit
EPRI Site 112 (oil-fired)
EPRI Site 118 (oil-fired)
Control Device
ESP, CS (inlet Hg concentration of 1.8
Hg/dscm)
ESP, CS (inlet Hg concentration of 1.4
ug/dscm)
Median
Mean
Standard deviation
Hg Removal %
83
41.7
62.35
6235
292
Reference
Camot, 1994b
Camot, 1994c
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 pm size range. These devices are used as gross paniculate removal
devices before ESPs or as APCDs on oil-fired units. Venturi scrubbers can be effective for paru'culate
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-AC22-93PC93J251. 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 # DE-AC22-93PC93251. 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.
Camot, 1994a. Preliminary draft emissions report for EPRI Site 115, Field Chemical Emissions
Monitoring Project. Prepared for Electric Power Research Institute. Carnot report No. EPRIE-
10106/R022C855.T. November 1994.
June 1996
B-ll
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Camot 1994b. Preliminary draft emissions report for EPRI Site 112, Field Chemical Emissions
Monitoring Project Prepared for Electric Power Research Institute. Carnot report No. EPREE-
10106/R016C374.T. March 1994.
Camot, 1994c. Preliminary draft emissions report for EPRI Site 118, Field Chemical Emissions
Monitoring Project Prepared for Electric Power Research Institute. Carnot report No. EPRJE-
10106/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 # DE-AC22-93PC93252. 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.
June 1996 B-12 SAB REVIEW DRAFT
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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, Reid 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, Reid 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, Reid 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.
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-ENV-
93-1049-7960. December 1993.
Southern Research Institute, 1993b. Preliminary draft emissions report for EPRI Site 110 (baseline
and with NOx control), Reid Chemical Emissions Monitoring Project. Prepared for Southern
Company Services. Report No. SRI-ENV-92-796-7496. October 1993.
June 1996 B-13 SAB REVIEW DRAFT
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Roy F. Western, 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 # 10316-011, 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 # 10316-011, Weston report #
DOE018G.RP1. December 1993.
June 1996 B-14 SAB REVIEW DRAFT
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APPENDIX C
EMISSION MODIFICATION FACTORS FOR
UTILITY BOILER EMISSION ESTIMATES
-------
-------
Table C-l
Emission Modification Factors for Utility Boiler Emission Estimates'*
Type of APCD or Boiler
Fabric Filter
Spray Dryer Adsorber (includes a fabric filter)
Electrostatic precipitator (cold-side)
Electrostatic precipitator (hot-side)
Electrostatic precipitator (oil-fired unit)
Paniculate matter scrubber
Fluidized gas desulfurization scrubber
Circulating fluidized bed combustor
Cyclone-fired boiler without NOx control (wet bottom, coal-fired)
Front-fired boiler without NOx control (dry bottom, coal-fired)
Front-fired boiler without NOx control (dry bottom, gas-fired)
Tangential-fired boiler without NOx control (before a hot-side ESP,
coal-fired)
Tangential-fired boiler with NOx control (before a hot-side ESP, coal-
fired)
Front-fired boiler without NOx control (dry bottom, oil-fired)
Front-fired boiler with NOx control (dry bottom, oil-fired)
Opposed-fired boiler without NOx control (dry bottom oil-fired)
Tangentially-fired boiler without NOx control (dry bottom, oil-fired)
TangentiaUy-fired boiler with NOx control (dry bottom, oil-fired)
Opposed-fired boiler with NOx control (dry bottom, coal-fired)
Front-fired boiler without NOx control (wet bottom, coal-fired)
Tangentially-fired boiler without NOx control (dry bottom, coal-fired)
Tangentially-fired boiler with NOx control (dry bottom, coal-fired)
Vertically-fired boiler with NOx control (dry bottom, coal-fired)
EMF Factor
0.626
0.701
0.684
1.000
0.315
0.957-
0.715
1.000
0.856
0.706
1.000
1.000
0.748
1.000
1.000
0.040
1.000
1.000
0.812
0.918
1.000
0.625
0.785
To calculate mercury control efficiency for a specific boiler/control device configuration, the EMF is subtracted from 1.
June 1996
C-l
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TECHNICAL REPORT DATA
(Please read Instructions on reverse before completing)
[.REPORT NO.
PA-452/R-96-001b
3. RECIPIENT'S ACCESSION NO
,E AND SUBTITLE
Mercury Study Report to Congress. Draft Submitted to U.S.
EPA's Science Advisory Board. Volume II. Inventory of
Anthropogenic Mercury Emissions in the United States.
5. REPORT DATE
1996
6 PERFORMING ORGANIZATION CODE
7. AUTHOR(S)
Ms. Martha H. Keating
8. PERFORMING ORGANIZATION REPORT NO.
9. PERFORMING ORGANIZATION NAME AND ADDRESS
U.S. Environmental Protection Agency
Office of Air Quality Planning and Standards
Research Triangle Park, NC 27711
10. PROGRAM ELEMENT NO.
11. CONTRACT/GRANT NO.
12. SPONSORING AGENCY NAME AND ADDRESS
Director
Office of Air Quality Planning and Standards
Office of Air and Radiation
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
13. TYPE OF REPORT AND PERIOD COVERED
Draft. June, 1996.
14. SPONSORING AGENCY CODE
EPA/200/04
15. SUPPLEMENTARY NOTES
U.S. EPA Project Officer: Martha H. Keating
ABSTRACT
is volume of the draft Mercury Study Report to Congress describes mercury emissions from anthropogenic
sources in the United States. The anthropogenic emissions summary estimates national mercury emissions
rates by source category for area and point sources including combustion, manufacturing and miscellaneous
point sources. Combustion point sources that dominate anthropogenic emissions are these: medical waste
incineration, municipal waste combustion, utility boilers, and commercial/industrial boilers. National
emission estimates are based on data from a 1990-1993 time-frame. Within the United States numerous
industrial and manufacturing processes contribute mercury emissions to the atmosphere. These emissions
occur nation-wide; however, the area east of the Mississippi River has the highest predicted concentrations of
emissions from anthropogenic sources in the U.S. The land areas having the greatest concentrations of
mercury emissions from anthropogenic sources of total mercury are the following: the urban corridor from
Washington DC to Boston, the Tampa and Miami areas of Florida, the larger urban areas of the Midwest and
Ohio Valley, and two sites in northern Texas. Sources of uncertainty and variability in these emissions
categories are described in this volume. Data needed to improve these estimates are described in this volume.
17.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
b. IDENTIFIERS/OPEN ENDED TERMS
c. COSATI Field/Group
cna
^PDI
I R
Mercury; Methylmercury; Air pollutants,
chemical; Clean Air Act; Control technology;
Indirect exposure; Bioaccumulate; Aquatic food
chain.
Air Pollution Control
^DISTRIBUTION STATEMENT
Release Unlimited
19. SECURITY CLASS (Repon)
Unclassified
21. NO. OF PAGES
196 pp.
20. SECURITY CLASS (Page)
Unclassified
22. PRICE
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