oEPA
United States
Environmental
Protection Agency
Report to Congress on Black Carbon
Department of the Interior, Environment, and Related Agencies
Appropriations Act, 2010
March 2012
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Printed on 100% recycled/recyclable process chlorine-free paper with 100% post-consumer fiber using vegetable oil-based ink.
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EPA-450/R-12-001
March 2012
Report to Congress on Black Carbon
U.S. Environmental Protection Agency
Office of Air Quality Planning and Standards
Office of Atmospheric Programs
Office of Radiation and Indoor Air
Office of Research and Development
Office of Transportation and Air Quality
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Acknowledgments
Lead Authors
Erika Sasser (Chair)
James Hemby (Vice-Chair)
Ken Adler
Susan Anenberg
Chad Bailey
Larry Brockman
Linda Chappell
Benjamin DeAngelo
Rich Damberg
Contributing Authors
Farhan Akhtar
Bryan Bloomer
Edmund Coe
Amanda Curry Brown
Penny Carey (retired)
Jason DeWees
Pat Dolwick
Robin Dunkins
Dale Evarts
Brian Gullett
John Guy (retired)
Beth Hassett-Sipple
Michael Hays
John Dawson
Neil Frank
Michael Geller
Gayle Hagler
Brooke Hemming
Lesley Jantarasami
Thomas Luben
John Mitchell
Jacob Moss
Carey Jang
Jim Jetter
Terry Keating
John Kinsey
Amy Lamson
Robin Langdon
Bill Linak
Bryan Manning
Allison Mayer
Harvey Michaels
Andy Miller
Ron Myers
Glenn Passavant
Venkatesh Rao
Joann Rice
Marcus Sarofim
Joseph Somers
Charlene Spells
Sara Terry
Matthew Witosky
Rob Pinder
Marc Pitchford
Adam Reff
Michael Rizzo
Charles Schenk
Darrell Sonntag
Larry Sorrels
Lauren Steele
Nicholas Swanson
Lori Tussey
Karen Wesson
Gil Wood
Rosa Yu
Additional Contributions to the Report
Jamie Bowers
Devin Hartman
Project Support
Lourdes Morales
Joseph Dougherty
Megan Melamed
Joseph Tikvart
Report Production
Sonoma Technology, Inc.
Steve Brown
Chelsea Jennings
Marina Michaels
Jana Schwartz
Editorial Support
Stratus Consulting
Nimmi Damodaran
Joe Donahue
Report to Congress on Black Carbon
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Acknowledgments
Peer Review
Council on Clean Air Compliance Analysis, Black Carbon Review Panel
C. Arden Pope, III (Chair) Ivan J. Fernandez Denise Mauzerall
Alberto Ayala H. Christopher Frey Surabi Menon
Michelle Bell Jan Fuglestvedt Richard L Poirot
Kevin J. Boyle D. Alan Hansen Armistead (Ted) Russell
Sylvia Brandt Joseph Helble Michael Walsh
Linda Bui MarkJacobson John Watson
James J. Corbett Jonathan Levy
EPA Science Advisory Board Staff
Stephanie Sanzone
Vanessa Vu
Contributing Federal Departments and Agencies
Centers for Disease Control and Prevention
Council on Environmental Quality
Department of Energy
Department of Transportation
Federal Highway Administration
Federal Aviation Administration
National Aeronautics and Space Administration
National Institute of Child Health and Human Development
National Institute of Environmental Health Sciences
National Institute of Standards and Technology
National Oceanic and Atmospheric Administration
Natural Resources Conservation Service
Office of Management and Budget
Office of Science and Technology Policy
United States Department of State
United States Agency for International Development
United States Department of Agriculture
United States Forest Service
Special Photo Credits
Cookstove in Guatemala (Cover): Nigel Bruce, University of Liverpool, UK
Brick Kiln in Kathmandu (Highlights): Sara Terry, U.S. Environmental Protection Agency
Outdoor Wood Boiler (Highlights): Philip Etter, Vermont Department of Environmental Conservation
Report to Congress on Black Carbon
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Highlights
Black carbon (BC) is the most strongly light-
absorbing component of particulate matter (PM),
and is formed by the incomplete combustion of
fossil fuels, biofuels, and biomass.
BC is emitted directly into the atmosphere in the
form of fine particles (PM2.5). The United States
contributes about 8% of the global emissions of
BC. Within the United States, BC is estimated to
account for approximately 12% of all direct PM2.5
emissions in 2005.
BC contributes to the adverse impacts on human
health, ecosystems, and visibility associated with
PM2.5.
BC influences climate by: 1) directly absorbing
light, 2) reducing the reflectivity ("albedo")
of snow and ice through deposition, and 3)
interacting with clouds.
The direct and snow/ice albedo effects of BC are
widely understood to lead to climate warming.
However, the globally averaged net climate effect
of BC also includes the effects associated with
cloud interactions, which are not well quantified
and may cause either warming or cooling.
Therefore, though most estimates indicate that BC
has a net warming influence, a net cooling effect
cannot be ruled out.
Sensitive regions such as the Arctic and the
Himalayas are particularly vulnerable to the
warming and melting effects of BC.
BC is emitted with other particles and gases,
many of which exert a cooling influence on
climate. Therefore, estimates of the net effect
of BC emissions sources on climate should
include the offsetting effects of these co-emitted
pollutants. This is particularly important for
evaluating mitigation options.
BC's short atmospheric lifetime (days to weeks),
combined with its strong warming potential,
means that targeted strategies to reduce BC
emissions can be expected to provide climate
benefits within the next several decades.
The different climate attributes of BC and
long-lived greenhouse gases make it difficult to
interpret comparisons of their relative climate
impacts based on common metrics.
Based on recent emissions inventories, the
majority of global BC emissions come from Asia,
Latin America, and Africa. Emissions patterns and
trends across regions, countries and sources vary
significantly.
Control technologies are available to reduce BC
emissions from a number of source categories.
BC mitigation strategies, which lead to reductions
in PM2.5, can provide substantial public health and
environmental benefits.
Considering the location and timing of emissions
and accounting for co-emissions will improve the
likelihood that mitigation strategies will be
Report to Congress on Black Carbon
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Highlights
properly guided by the balance of climate and
public health objectives.
Achieving further BC reductions, both
domestically and globally, will require adding a
specific focus on reducing direct PM2.5 emissions
to overarching fine particle control programs.
The most promising mitigation options identified
in this report for reducing BC (and related
"soot") emissions are consistent with control
opportunities emphasized in other recent
assessments.
- United States: The United States will achieve
substantial BC emissions reductions by 2030,
largely due to controls on new mobile diesel
engines. Other source categories in the United
States, including stationary sources, residential
wood combustion, and open biomass burning
also offer potential opportunities.
- Global: The most important BC emissions
reduction opportunities globally include
residential cookstoves in all regions; brick kilns
and coke ovens in Asia; and mobile diesels in
all regions.
- Sensitive Regions: To address impacts in
the Arctic, other assessments have identified
the transportation sector; residential heating;
and forest, grassland and agricultural
burning as primary mitigation opportunities.
In the Himalayas, studies have focused
on residential cooking; industrial sources;
and transportation, primarily on-road and
off-road diesel engines.
A variety of other options may also be suitable
and cost-effective for reducing BC emissions,
but these can only be identified with a tailored
assessment that accounts for individual countries'
resources and needs.
Despite some remaining uncertainties about BC
that require further research, currently available
scientific and technical information provides
a strong foundation for making mitigation
decisions to achieve lasting benefits for public
health, the environment, and climate.
iv
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Table of Contents
Acknowledgments i
Highlights iii
List of Figures xii
List of Tables xxi
List of Acronyms xxiii
Glossary xxviii
Executive Summary 1
Chapter 1 Introduction 11
1.1 Key Questions Addressed in this Report 12
1.2 Other Recent Assessments of BC 13
1.3 Organization of this Report 14
Chapter 2 Black Carbon and Its Effects on Climate 17
2.1 Summary of Key Messages 17
2.2 Introduction 18
2.3 Defining Black Carbon and Other Light-Absorbing PM 20
2.4 Key Attributes of BC and Comparisons to GHGs 25
2.5 The Role of Co-Emitted Pollutants and Atmospheric Processing 28
2.6 Global and Regional Climate Effects of Black Carbon 32
2.6.1 Global and Regional Radiative Forcing Effects of BC: Overview 33
2.6.2 Impact of BC Radiative Forcing on Temperature and Melting of Ice and Snow . . . 47
2.6.3 Other Impacts of BC 49
2.6.4 BC Impacts in the Arctic 52
2.6.5 BC Impacts in the Himalayas 55
2.6.6 Summary of BC Impacts in Key Regions 56
2.7 Metrics for Comparing Black Carbon Impacts to Impacts of Other Climate Forcers. . . 56
2.7.1 Metrics along the Cause and Effect Chain 57
2.7.2 Commonly-Used Metrics for GHGs 59
2.7.3 Applicability of Climate Metrics to BC 61
2.7.4 Using Metrics in the Context of Climate Policy Decisions 64
2.8 Key Gaps in Understanding and Expressing the Climate Impacts of BC 66
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Chapter 3 Black Carbon Effects on Public Health and the Environment 67
3.1 Summary of Key Messages 67
3.2 Introduction 67
3.3 Health Effects Associated with BC 67
3.3.1 Key Health Endpoints Associated with Exposure to PM 67
3.3.2 Health Effects Related to Ambient BC Concentrations 69
3.3.3 Health Effects Related to Indoor BC Exposures 81
3.4 Non-Climate Welfare Effects of PMZ5, Including BC 82
3.4.1 Role of BC in Visibility Impairment 82
3.4.2 Role of BC in Crop Damage and Other Environmental Impacts 83
3.5 Key Uncertainties Regarding Health/Environmental Impacts of BC 83
Chapter 4 Emissions of Black Carbon 85
4.1 Summary of Key Messages 85
4.2 Introduction 85
4.3 U.S. Black Carbon Emissions 86
4.3.1 Summary of Emissions Methodology 86
4.3.2 U.S. Black Carbon Emissions: Overview and by Source Category 87
4.4 Global Black Carbon Emissions 95
4.4.1 Summary of Global Black Carbon Emissions by Region and Source Category. ... 96
4.4.2 Black Carbon Emissions North of the 40th Parallel 104
4.4.3 Alternative Estimates of Global and Regional Emissions 106
4.4.4 Inventory Comparisons for U.S. Black Carbon Emissions 107
4.5 Long-Range Transport of Emissions 109
4.6 Historical Trends in Black Carbon Emissions 112
4.6.1 U.S. Black Carbon Emissions Trends 112
4.6.2 Global Black Carbon Emissions Trends 112
Chapter 5 Observational Data for Black Carbon 115
5.1 Summary of Key Messages 115
5.2 BC and Other Light-Absorbing Carbon: Measurement Methods 116
5.3 Ambient Concentrations of BC 118
5.3.1 Major Ambient Monitoring Networks 118
5.3.2 Global Ambient Concentrations 119
5.3.3 Comparison of Urban and Rural Concentrations Globally 119
5.3.4 BC as a Percentage of Measured Ambient PM2.5 Concentrations
in the United States 123
5.4 Trends in Ambient BC Concentrations 125
5.4.1 Trends in Ambient BC Concentrations in the United States
and the United Kingdom 125
5.4.2 Trends in Ambient BC Concentrations in the Arctic 128
5.5 Remote Sensing Observations 129
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5.6 BC Observations from Surface Snow, Ice Cores, and Sediments 132
5.6.1 Measurement Approach 132
5.6.2 Surface Snow Data 132
5.6.3 Ice Core Data 134
5.6.4 Sediment Data 135
5.6.5 Arctic BC Snow and Ice Data - Source Identification 136
5.7 Limitations and Gaps in Current Ambient Data and Monitoring Networks 137
Chapter 6 Benefits of Reducing Black Carbon Emissions 139
6.1 Summary of Key Messages 139
6.2 Introduction 140
6.3 Public Health Benefits of Reducing Black Carbon Emissions 140
6.3.1 Health Benefits in the United States 140
6.3.2 Global Health Benefits 144
6.3.3 Conclusions Regarding Potential Health Benefits 148
6.4 Climate Benefits of Reducing Black Carbon Emissions 148
6.4.1 Studies Estimating Physical Climate Benefits 149
6.4.2 Comparing Climate Benefits of Reductions in BC vs. CO2 154
6.4.3 Valuing the Climate Benefits of BC Mitigation 156
6.4.4 Conclusions Regarding Climate Benefits 158
6.5 Environmental Benefits of BC Reductions 159
6.5.1 Visibility Impacts 159
6.5.2 Ecosystem Impacts 159
6.5.3 Materials Co-benefits 160
6.5.4 Conclusions Regarding Environmental Benefits 160
6.6 Conclusions 160
Chapter 7 Mitigation Overview: Designing Strategies for Public Health
and Near-Term Climate Protection 161
7.1 Summary of Key Messages 161
7.2 Introduction 162
7.3 Effect of Existing Control Programs 162
7.4 Future Black Carbon Emissions 163
7.5 Key Factors to Consider in Pursuing BC Emissions Reductions 166
7.5.1 Defining Goals: Climate, Health and Environmental Outcomes 167
7.5.2 Identifying Opportunities for Emissions Reductions 168
7.5.3 Key Considerations 169
7.6 Applying the Mitigation Framework 172
7.7 Conclusions 173
Chapter 8 Mitigation Approaches for Mobile Sources 175
8.1 Summary of Key Messages 175
8.2 Introduction 176
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8.3 Emissions Trajectories for Mobile Sources 176
8.4 New Engine Standards in the United States 177
8.4.1 On-road and Nonroad Diesel Engines 180
8.4.2 On-road and Nonroad Gasoline Engines 182
8.4.3 Other Mobile Sources - Commercial Marine Vessels, Locomotives, and Aircraft . 183
8.5 New Engine Standards Internationally 184
8.5.1 International Regulations of Diesel Fuel Sulfur Levels 184
8.5.2 Standards for New Engines Outside the United States 185
8.6 Mitigation Approaches for In-use Mobile Sources in the United States 186
8.6.1 Available Retrofit Technologies and Strategies for In-use Engines 187
8.6.2 Cost-Effectiveness of Retrofits 190
8.6.3 Applicability of Diesel Retrofits 191
8.6.4 Experience with Diesel Emissions Reduction Programs in the United States . . . 191
8.7 Mitigation Approaches for In-use Mobile Engines Internationally 194
Chapter 9 Mitigation Approaches for Stationary Sources 195
9.1 Summary of Key Messages 195
9.2 Introduction 195
9.3 Emissions from Key Stationary Source Categories 196
9.4 Available Control Technologies for Stationary Sources 197
9.4.1 Fabric Filters 197
9.4.2 Electrostatic Precipitators 198
9.4.3 Diesel Particulate Filters and Oxidation Catalysts 198
9.5 Cost-Effectiveness of PM Control Technologies 199
9.6 Mitigation Approaches Other than PM Control Technologies 200
9.6.1 Process Modification/Optimization 200
9.6.2 Fuel Substitution and Source Reduction Approaches for PM 200
9.7 Mitigation Approaches for Stationary Sources Internationally 201
9.7.1 Brick Kilns 201
9.7.2 Coke Production/Iron and Steel Production 202
9.7.3 Power Generation and Industrial Boilers 202
9.7.4 Oil and Gas Flaring 203
9.8 Technical and Research Needs 203
Chapter 10 Mitigation Approaches for Residential Heating and Cooking 205
10.1 Summary of Key Messages 205
10.2 Introduction 206
10.3 Residential Wood Combustion in Developed Countries 207
10.3.1 Emissions from Residential Wood Combustion 207
10.3.2 Approaches for Controlling Emissions from RWC 207
10.3.3 Emissions Standards for New Wood-burning Units 208
10.3.4 Mitigation Opportunities for In-Use RWC Sources 209
10.3.5 Additional Regulatory Approaches to Limiting Wood Smoke Emissions 211
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10.3.6 Wood Smoke Reduction Resource Guide 212
10.4 Residential Cookstoves in Developing Countries 213
10.4.1 Emissions from Cookstoves 214
10.4.2 Technologies and Approaches for Controlling Emissions from Cookstoves . . . 214
10.4.3 Programmatic Considerations for Cookstove Mitigation 221
Chapter 11 Mitigation Approaches for Open Biomass Burning 227
11.1 Summary of Key Messages 227
11.2 Introduction 227
11.3 Emissions from Open Biomass Burning 227
11.4 Fire as a Resource Management Tool 229
11.5 Smoke Mitigation Technologies and Approaches in the United States 230
11.5.1 Managing Smoke 231
11.5.2 Fire Prevention Techniques 233
11.6 Mitigation Technologies and Approaches Globally 234
Chapter 12 Key Black Carbon Mitigation Opportunities and
Areas for Further Research 237
12.1 Summary of Key Messages 237
12.2 Introduction 238
12.3 Controlling Black Carbon as Part of Broader PM2.5 Mitigation Program 239
12.4 Key Black Carbon Mitigation Opportunities 240
12.4.1 U.S. Black Carbon Mitigation Opportunities 241
12.4.2 Global Black Carbon Mitigation Opportunities 242
12.4.3 Other Mitigation Options 245
12.5 Key Policy-Relevant Scientific Uncertainties 245
12.6 High Priority Research Needs for Black Carbon 247
Appendix 1 Ambient and Emissions Measurement of Black Carbon 251
Appendix 2 Black Carbon Emissions Inventory Methods and Comparisons 271
Appendix 3 Studies Estimating Global and Regional Health Benefits of
Reductions in Black Carbon 285
Appendix 4 Efforts to Limit Diesel Fuel Sulfur Levels 287
Appendix 5 U.S. Emission Standards for Mobile Sources 295
Appendix 6 International Emission Standards for Heavy-Duty Vehicles 299
Appendix 7 Research Needs 303
Bibliography 311
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List of Figures
Figure A. BC Emissions by Major Source Category 2
Figure B. Regional Variability in Direct Radiative Forcing and Snow/Ice Albedo Forcing for
BC from All Sources 3
Figure C. BC Emissions, 2000, Gg 5
Figure D. Policy Framework for Black Carbon Mitigation Decisions 7
Figure 2-1. Effects of BC on Climate, as Compared to GHGs 19
Figure 2-2. BC Images 20
Figure 2-3. Representative Examples of Filter Samples Collected from Different Sources 21
Figure 2-4. Light Absorption by BC, BrC, and Ambient Mixtures 22
Figure 2-5. TEM Image of a BrC Particle 24
Figure 2-6. Coarse Urban PM (Diameter > 2.5 microns) with a Black Surface Coating 24
Figure 2-7. Climate Response to Emissions of Pollutants with Different Lifetimes 28
Figure 2-8. Projected Global Mean Temperatures under Various Scenarios Relative to the
1890-1910 Average 29
Figure 2-9. Particle Transformation in the Atmosphere, from Point of Emission to Deposition. ... 30
Figure 2-10. Components of Global Average Radiative Forcing for Emissions of Principal Gases,
Aerosols, and Aerosol Precursors, based on IPCC estimates 34
Figure 2-11. Estimates of Radiative Forcing from BC Emissions Only 35
Figure 2-12. Estimates of Direct Radiative Forcing from BC Emissions Only 37
Figure 2-13. Direct Radiative Forcing (W m 2) of BC from All Sources 38
Figure 2-14. Estimates of Snow and Ice Albedo Radiative Forcing Effects from BC Emissions Only. . 41
Figure 2-15. Snow and Ice Albedo Forcing by BC 42
Figure 2-16. Estimates of Direct Radiative Forcing from OC Emissions Only 43
Figure 2-17. Estimates of Direct Radiative Forcing from BC and OC Emissions 43
Figure 2-18. Direct Forcing by OC from All Sources 44
Figure 2-19. Global Radiative Forcing Due to Perpetual Constant Year 2000 Emissions, Grouped
By Sector, at (a) 2020 and (b) 2100 and Showing the Contribution from Each Species. . 45
Figure 2-20. Spatial Distribution of Change in Mean Snow Water Equivalent (SWE, mm) for March. . 48
Figure 2-21. BC Concentrations in the ZD Glacier 49
Figure 2-22. Surface Dimming by Anthropogenic Aerosols (W m2) 51
Figure 2-23. Evidence of Arctic Ice Melt 53
Figure 2-24. Cause and Effect Chain from Emissions to Climate Change, Impacts, and Damages. . . 57
Figure 2-25. Ranges and Point Estimates for Regional Estimates of GWP Values for One-Year
Pulse Emissions of BC for Different Time Horizons 62
Figure 2-26. Ranges and Point Estimates for Regional Estimates of GTP Values for One-Year
Pulse Emissions of BC for Different Time Horizons. . 63
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List of figures
Figure 3-1. Conceptual Diagram of the Epidemiological Evidence for the Association of BC
with the Continuum of Cardiovascular Effects 76
Figure 4-1. BC and OC Fractions of PM2.5 Emissions for the Highest BC Emitting Non-Mobile
Source Categories in the United States 86
Figure 4-2. Heavy-Duty Diesel PM2.5 Emissions Profile 87
Figure 4-3. Contribution to Primary PM2.5, BC, and OC Emissions by Mega Source Categories. ... 88
Figure 4-4. U.S. BC Emissions (tons) for Major Source Categories 89
Figure 4-5. U.S. BC Emissions from all Mobile Source Categories 92
Figure 4-6. U.S. BC Emissions from all Biomass Combustion Source Categories
(250,000 short tons) 94
Figure 4-7. Acres Burned per Year in Alaskan Wildfires, 2002-2010 95
Figure 4-8. Global BC Emissions based on Year 2000 Estimates, in Gigagrams (Gg) 97
Figure 4-9. BC Emissions (Aggregate) by Selected World Region, 2000 (Gg) 97
Figure 4-10. Global Distribution of BC and OC Emissions by Major Source Category 102
Figure 4-11. BC Emissions by World Region, 2000 (Gg) 103
Figure 4-12. Global BC Emissions by Source Categories and Region 104
Figure 4-13. Geographical Distribution of Global BC Emissions by Latitude 105
Figure 4-14. Comparison of Regional Inventories for China, India, and Indonesia with
AR5 Estimates 107
Figure 4-15. Relative Importance of Different Regions to Annual Mean Arctic BC
Concentrations at the Surface and in the Upper Troposphere (250 hPa) 110
Figure 4-16. Potential for Transport of U.S. Emissions to the Arctic Ill
Figure 4-17. Historical Growth in Emissions of BC and OC, Segregated by Fuel and World Region. 113
Figure 4-18. Historical Reconstruction of Global Emissions Trends 113
Figure 4-19. BC Emissions (Tg /y) in the United States, United Kingdom, and China 114
Figure 5-1. Measurement of the Carbonaceous Components of Particles 117
Figure 5-2. Ambient BC Measurement Locations Worldwide 119
Figure 5-3. Spatial Distribution of Global BC Data 121
Figure 5-4. Annual Mean BC Concentrations (u.g m3) for 2005-2008 in the United States 122
Figure 5-5. Urban BC Gradients for New York City 123
Figure 5-6. Composition of PM2.5 for 15 Selected Urban Areas in the United States 124
Figure 5-7a. Trends in Black Smoke Measurements (ug/m3) in the United Kingdom, 1954-2005. . . 124
Figure 5-7b. Comparison of Ambient Black Smoke Measurements (ug/m3, annual average) with
Estimated BC emissions (Tg) in the United Kingdom, 1955-2000 124
Figure 5-8. Ambient BC Trends in Washington, D.C 125
Figure 5-9. Trends in BC at All IMPROVE Network Stations with Sufficient Data between 1
March 1990 and 29 February 2004 126
Figure 5-10. Estimated Annual Average Ambient BC Concentrations in the San Francisco Bay
Area vs. Diesel Fuel Consumption 127
Figure 5-11. Ambient BC Trends in Boston (Harvard School of Public Health location) 127
Figure 5-12. Ambient BC Trends (2002-2010) in the United States 128
Figure 5-13. The Annual Mean BC Concentrations Measured at Alert (a), Barrow (b), and
Zeppelin (c) and Split into Contributions from the Four Transport Clusters 129
Figure 5-14. Aerosol Absorption Optical Depth (AAOD) from AERONET (1996-2006) and
OMI (2005-2007) 130
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List of Figures
Figure 5-15. Absorption Angstrom Exponent (AAE) Values for AAOD Spectra Derived from
AERONET Data 130
Figure 5-16. AERONET AOD and AAOD as a Percentage of AOD 131
Figure 5-17. Locations of BC Measurements in Surface Snow and Shallow Snow Pits 133
Figure 5-18. BC Concentrations in Surface Snow in Arctic and Subarctic Areas of the Northern
Hemisphere 133
Figure 5-19. BC Ice Core Records Worldwide Labeled by Their Identifying Name 134
Figure 5-20. Atmospheric BC determined by Husain et al. (2008), for the Adirondack Region
from 1835 to 2005 135
Figure 5-21. Annual Average Concentrations of (a) BC and VA and (b) BC and Non-Sea-Salt
Sulfur (nss-S) 136
Figure 5-22. Sources of BC in Arctic Snow 137
Figure 6-1. Estimated Global Mortality Benefits of Black Carbon Reductions 145
Figure 6-2. Annual Avoided Premature Cardiopulmonary and Lung Cancer Deaths Per Unit BC
Emissions Reduced versus Total BC Emissions (Gg) for Particular Source Sectors
within Each Region 146
Figure 6-3. Comparison of Premature Mortality by Region (millions of premature deaths
annually) 147
Figure 6-4. Global Radiative Forcing Due to Perpetual Constant Year 2000 Emissions,
Grouped by Sector, in 2020 and 2100 150
Figure 6-5. Observed Deviation of Temperature to 2009 and Projections under
Various Scenarios 151
Figure 6-6. Summary of Normalized Net Forcing per Unit of Emissions 152
Figure 6-7. Contribution to Radiative Forcing of Carbonaceous Aerosol Emissions within
Different Latitude Bands 153
Figure 6-8. Integrated Forcing by Aerosols Emitted from Burning 1 Kg of Fuel from Different
Sources 155
Figure 6-9. Cause and Effect Chains for (a) CO2 and (b) BC from Emissions to Damages 157
Figure 7-1. Global BC Emissions Forecasts for Various Sectors under Alternative IPCC SRES
Scenarios (in teragrams (Tg) of carbon) 164
Figure 7-2. Black Carbon Emissions Growth, 2000-2030 under IPCC A1B Scenario 165
Figure 7-3. Future Emissions of BC under IPCC Representative Concentration Pathways, 2000-
2050 (Gg/year) 165
Figure 7-4. Policy Framework for Black Carbon Mitigation Decisions 167
Figure 7-5. OC (left) and BC (right) Emissions from Key U.S. and Global Emissions Source
Categories, Expressed as a Fraction of Total Carbon (OC + BC) Emissions from that
Category 170
Figure 8-1. Estimated Changes in Emissions of (a) BC, (b) OC, and (c) Direct PM25 from Mobile
Sources in the United States, 1990-2030 179
Figure 10-1. OC/BC Emission Ratios by Source Category and Fuel Type 208
Figure 10-2. Cost Per Ton PM2.5 Reduced for Replacing Non-EPA-Certified Wood Stove with
EPA-Certified Woodstove (in 2010$) 210
Figure 10-3. Cost Per Ton PM25 Reduced ($/Ton) for the Addition of an Insert into a Fireplace
(2010$) 211
Figure 10-4. The Turbococina Stove 215
Figure 10-5. Woman Prepares Banku on a BioLite HomeStove in Kintampo, Ghana 215
Figure 10-6. CleanCook Ethanol Stove 216
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List of figures
Figure 10-7. Charcoal Stoves 217
Figure 10-8. Philips Woodstove (forced draft) Manufactured in Lesotho 218
Figure 10-9. Rocket Stoves 219
Figure 10-10. Prakti Double-Pot Woodstove with Chimney 220
Figure 10-11. Number of Improved Stoves Sold by PCIA Partners, 2003-2010 222
Figure 10-12. Potential Growth in the Number of Households Adopting Clean Cookstoves
Globally through 2020 224
Figure 12-1. Key Policy-Relevant Scientific Uncertainties Related to BC 246
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List of Tables
Table 2-1. Comparison of BC to CO2 on the Basis of Key Properties that Influence the Climate. 26
Table 2-2. Examples of Particle Types and Mixtures Present in Combustion Plumes 31
Table 2-3. Summary of UNEP/WMO Assessment Estimates of Radiative Forcing Effects of BC. . 36
Table 2-4. Overview of the Different Aerosol Cloud Effects 39
Table 2-5. Overview of the Different Aerosol Indirect Effects and Their Implications for
Global Dimming and Precipitation 50
Table 2-6. Arctic Temperature Impacts from Emissions of BC from Different Sectors 54
Table 2-7. Climate Effects of BC in the United States, Asia, and the Arctic (Summary) 56
Table 2-8. Examples of Commonly Used Metrics for GHGs 58
Table 3-1. Summary of Causal Determinations for Exposure to PM2.5 from 2009 PM ISA 68
Table 3-2. Summary of Epidemiological Studies of BC and Cardiovascular Health Outcomes. . 70
Table 3-3. Summary of Epidemiological Studies of BC and Respiratory Health Outcomes. ... 77
Table 3-4. Summary of Epidemiological Studies of BC and Mortality 79
Table 4-1. 2005 U.S. Emissions (tons) and Ratios of Emissions by Mega Source Category. ... 89
Table 4-2. U.S. Emissions of PM25, BC, and OC (short tons) 90
Table 4-3. National Level U.S. Emissions of PM2.5, BC, and OC for Biomass Combustion
Sources in 2002/2005 (short tons) 94
Table 4-4. Global BC Emissions in 2000 (in Gg). Transport Includes Aircraft and Shipping. ... 98
Table 4-5. Global OC Emissions in 2000 (in Gg). Transport Includes Only Aircraft 100
Table 4-6. OC/BC Ratios by Broad Source Categories 103
Table 4-7. A Comparison of BC Emissions Nationally to Those from Sources "North of 40th
Parallel" in 2005 (short tons) 106
Table 4-8. Comparison of BC and OC Emissions (in Gg) for the United States between AR5
Global Inventories and EPA Inventories 108
Table 5-1. Description of BC Measurement Techniques 116
Table 5-2. Summary of Selected Global BC Ambient Concentrations for Urban and Rural/
Remote Areas 120
Table 6-1. PM2.5 Health Endpoints Included in EPA's Regulatory Impact Analyses 141
Table 6-2. Changes in Key Health Effects Outcomes in the United States Associated with
PM2.5 Resulting from the 1990 CAA Amendments 142
Table 6-3. List of Benefits, Costs, and Benefit to Cost Ratios for U.S. Rules with Direct PM
Reductions (Billions 2010$) 142
Table 6-4. Direct PM25 National Average Benefits per Ton Estimates by Source Category for
the United States (3% Discount Rate, Thousands of 2010$) 143
Table 8-1. Mobile Source BC, OC, and PM2.5 Emissions 1990-2030 (short tons) 178
Table 8-2. Cost Estimates for Particulate Matter Controls on New Diesel Engines (2010$),
based on Recent U.S. EPA rulemakings 182
Table 9-1. PM Control Costs for ICI Boilers 200
Table 11-1. Types of Open Biomass Burning 228
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List of Acronyms
AAE Absorption Angstrom Exponent
AAOD Aerosol Absorption Optical
Depth
ABC Atmospheric Brown Cloud
ACIA Arctic Climate Impact
Assessment
ACS American Cancer Society
AERONET Aerosol Robotic (Measurement)
Network
AIE Aerosol Indirect Effect
AMAP Arctic Monitoring and
Assessment Program
AOD Aerosol Optical Depth
AP-42 Compilation of Air Pollutant
Emission Factors (EPA)
APU Auxiliary Power Unit
AQS Air Quality System (EPA)
AR5 Fifth Assessment Report of the
Intergovernmental Panel on
Climate Change
ARIES Aerosol Research Inhalation
Epidemiology Study
BB Biomass Burning
BC Black Carbon
BCa Apparent Black Carbon
BF Biofuel
BrC Brown Carbon
BS Black Smoke
BSG Biofuel Soot and Gases
Cl Commercial Marine Engines less
than 5 liters/cylinder
C2 Commercial Marine Engines
between 5-30 liters/cylinder
C3 Commercial Marine Engines
greater than 30 liters/cylinder
CAA Clean Air Act
CALIPSO Cloud-Aerosol Lidar and
Infrared Pathfinder Satellite
Observation
CAPMoN Canadian Air and Precipitation
Monitoring Network
CARB California Air Resources Board
CAWNET China Atmosphere Watch
Network
CCN Cloud Condensation Nuclei
CCVS Closed Crankcase Ventilation
System
CF4 Carbon Tetrafluoride
CH4 Methane
CI Compression Ignition
CLRTAP Convention on Long-Range
Transboundary Air Pollution
CMAQ Congestion Mitigation and Air
Quality
CNG Compressed Natural Gas
CO Carbon Monoxide
CO2 Carbon Dioxide
COM Coefficient of Haze
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List of Acronyms
CRF Concentration-Response
Function
CSN Chemical Speciation Network
DALYs Disability Adjusted Life Years
DERA Diesel Emissions Reduction Act
DOC Diesel Oxidation Catalyst
DPF Diesel Particulate Filter
EC Elemental Carbon
ECa Apparent Elemental Carbon
ECA Emissions Control Area
EDX Energy Dispersive X-ray
Spectroscopy
EF Emission Factor
El Emission Index
EMEP European Monitoring and
Evaluation Programme
EPA Environmental Protection
Agency
ER Emission Reduction Factor
ERT Emission Reduction Technique
ESP Electrostatic Precipitator
ESRL/GMD Earth System Research
Laboratory/Global Monitoring
Division (NOAA)
EU European Union
EURO European Emissions Standards
EUSAAR European Supersites for
Atmospheric Aerosol Research
FAO Food and Agriculture
Organization
FF Fossil Fuel
FHWA Federal Highway Administration
FRM Federal Reference Method
g/bhp-hr
GACC
GAW Aerosol
GCP
GDI
GDP
GSFC
Gg
GHG
GIZ
GLAS
GTP
GWP
H2O
HAP
HC
HFC
HHK
HKHT
HULIS
hPa
IAM
ICI
IEA
Grams per Break Horsepower-
Hour
Global Alliance for Clean
Cookstoves
Global Atmospheric Watch
Aerosol Program
Global Cost Potential
Gasoline Direct Injection
Global Damage Potential
Goddard Space Flight Center
Gigagram (109 g = 1 kilotonne)
Greenhouse Gas
German Agency for
International Development
(Deutsche Gesellschaft fur
Internationale Zusammenarbeit)
Geoscience Laser Altimeter
System (NASA)
Global Temperature Potential
Global Warming Potential
Water
Hazardous Air Pollutant
Hydrocarbon
Hydrofluorocarbon
Hybrid Hoffman Kiln
Hindu-Kush Himalayan Tibetan
Region
Humic-Like Substances
Hectopascal (a unit of
barometric pressure)
Integrated Assessment Model
Industrial, Commercial and
Institutional (Boilers)
International Energy Agency
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List of Acronyms
IGAC International Global
Atmospheric Chemistry Project
IIASA International Institute for
Applied Systems Analysis
IMPROVE Interagency Monitoring of
Protected Visual Environments
IPCC Intergovernmental Panel on
Climate Change
IR Infrared
ISA Integrated Science Assessment
LAC Light-Absorbing Carbon
LDGV Light Duty Gasoline Vehicle
LEV Low Emissions Vehicle
LII Laser Induced Incandescence
LNG Liquefied Natural Gas
LPG Liquid Petroleum Gas
LPM Liters per Minute
LTO Landing and Take-Off
MAC Mass Absorption Coefficients
MACT Maximum Achievable Control
Technology
MANE-VU Mid-Atlantic/Northeast
Visibility Union
MARAMA Mid-Atlantic Regional Air
Management Association
MARPOL International Convention on the
Prevention of Pollution from
Ships
MI Myocardial Infarction
MISR Multi-angle Imaging
Spectroradiometer
Mm Millimeter
MODIS Moderate Resolution Imaging
Spectroradiometer
MOVES Mobile Vehicle Emission
Simulator Model (EPA)
N2O Nitrous Oxide
NAAQS National Ambient Air Quality
Standard
NAPS National Air Pollution
Surveillance Network
NATA National Air Toxics Assessment
NCDC National Clean Diesel Campaign
NCO-P Nepal Climate Observatory
Pyramid
NEI National Emissions Inventory
(EPA)
NESCAUM Northeast States for
Coordinated Air Use
Management
NESHAP National Emissions Standards
for Hazardous Air Pollutants
NGO Non-Governmental
Organization
NIOSH National Institute for
Occupational Safety and Health
NMIM National Mobile Inventory
Model
NOAA National Oceanic and
Atmospheric Administration
NOX Nitrogen Oxides
NRC National Research Council
NSPS New Source Performance
Standard
Nss-S Non-Sea-Salt Sulfur
O3 Ozone
OC Organic Carbon
OCa Apparent Organic Carbon
OC/BC Organic Carbon to Black Carbon
Ratio (also OCBC)
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List of Acronyms
OECD
OM
OMI
PAH
PBL
PCFV
PCIA
PDPF
PIC
PM
POM
RCP
RHR
RICE
RM
RPO
RWC
SAGE
SCC
SCR
SEARCH
SEM
SFP
SLCF
Organisation for Economic Co-
operation and Development
Organic Matter
Ozone Monitoring Instrument
Polycyclic Aromatic
Hydrocarbon
Planetary Boundary Layer
Partnership for Clean Fuels and
Vehicles
Partnership for Clean Indoor Air
Partial Diesel Particulate Filters
Product of Incomplete
Combustion
Particulate Matter (for related
terms, such as PM2.s and PMi0;
see Glossary)
Polycyclic Organic Matter
Representative Concentration
Pathway
Regional Haze Rule
Reciprocating Internal
Combustion Engine
Raman Microspectroscopy
Regional Planning Organization
Residential Wood Combustion
System for Assessing Aviation's
Global Emissions (FAA)
Social Cost of Carbon
Selective Catalytic Reduction
Southeastern Aerosol Research
and Characterization
Scanning Electron Microscopy
Specific Forcing Pulse
Short-Lived Climate Forcer
SNV
S02
SOA
SPARC
SRES
STN
STRE
TEM
TERP
TF HTAP
Tg
TOA
TOMS
TOR
TOT
jjg/m3
ULSD
UNEP
UNFCCC
USAID
USDA
USGCRP
Netherlands Development
Organization
Sulfur Dioxide
Secondary Organic Aerosols
Stratospheric Processes and
their Role in Climate (World
Climate Research Programme)
Special Report on Emissions
Scenarios (IPCC)
Speciation Trends Network
Surface Temperature Response
per unit continuous Emission
Transmission Electron
Microscopy
Texas Emissions Reduction Plan
Task Force on Hemispheric
Transport of Air Pollution
Teragram (1012 g = 1
megatonne)
Top of the Atmosphere
Total Ozone Mapping
Spectrometer
Thermal/Optical Reflectance
Thermal/Optical Transmittance
Micrograms per cubic meter
Ultra-Low-Sulfur Diesel
United Nations Environment
Programme
United Nations Framework
Convention on Climate Change
United States Agency for
International Development
United States Department of
Agriculture
United States Global Climate
Change Research Program
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List of Acronyms
UV
VA
VMT
VOC
VSBK
VSL
Wm2
Ultraviolet
Vanillic Acid
Vehicle Miles Traveled
Volatile Organic Compound
Vertical Shaft Brick Kiln
Value of a Statistical Life
Watts per square meter (also
W/m2)
WAIS
WHO
WMO
WRAP
WTP
West Antarctic Ice Sheet
World Health Organization
World Meteorological
Organization
Western Regional Air
Partnership
Willingness to Pay
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Glossary
Aerosol
A mixture of gases and suspended solid and/or liquid particles, with a typical size between 0.01
and 10 micrometers and residing in the atmosphere for at least several hours. Aerosols may be of
either natural or anthropogenic origin. Often the term is used interchangeably with "particle" or
"particulate matter."
Aerosol Absorption Optical Depth
A quantitative measure of light extinction within a vertical column of atmosphere due to absorption
by aerosols.
Aerosol Optical Depth
A quantitative measure of light extinction within a vertical column of atmosphere due to aerosol
absorption or scattering. Pollution and cloud-free portions of the atmosphere have a low aerosol
optical depth, while highly polluted or densely cloudy skies have a high optical depth.
Aging
The changes that occur to a particle over the course of its atmospheric lifetime, including changes in
size or chemical composition.
Agricultural Burning
The planned burning of vegetative debris from agricultural operations; or, the use of fire as a
method of clearing land for agricultural use or pastureland.
Albedo
The fraction of solar radiation reflected by a surface or object, often expressed as a percentage.
Light-colored surfaces (such as those covered by snow and ice) have a high albedo; dark surfaces
(such as dark soils, vegetation and oceans) have a low albedo.
Arctic Haze
A persistent reddish-brown haze visible in the atmosphere at high latitudes in the Arctic due to air
pollution, including black carbon, organic carbon, and sulfate particles.
Atmospheric Brown Clouds
Pollution clouds consisting of combinations of black carbon, brown carbon, sulfates, organics, dust,
and other components. Atmospheric brown clouds are more common in Asia, southern Africa, and
the Amazon Basin. They have been linked to surface dimming and a decrease in vertical mixing
(which exacerbates air pollution episodes), and they contribute to changes in the pattern and
intensity of rainfall (particularly with respect to monsoon circulation in South Asia).
Atmospheric Lifetime
The approximate amount of time it would take for the atmospheric concentration of a pollutant to
return to its natural level (assuming emissions cease) as a result of either being converted to another
chemical compound or being taken out of the atmosphere via a sink. This time depends on the
pollutant's sources and sinks as well as its reactivity. Average lifetimes for air pollutants can vary
from days to weeks (black carbon and ozone) to more than a century (e.g., chlorofluorocarbons and
carbon dioxide).
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Glossary
Atmospheric Residence Time
See atmospheric lifetime.
Atmospheric Transport
The movement of chemical species through the atmosphere as a result of large-scale atmospheric
motions. Transport distances are a function of atmospheric lifetimes, emission location, and overall
meteorological activity.
Baghouse
See fabric filter.
Biofuels
Biofuels are non-fossil carbon-based fuels derived from organic materials (biomass), including plant
materials and animal waste.
Biomass
In the context of energy, the term biomass is often used to refer to organic materials, such as wood
and agricultural wastes, which can be burned to produce energy or converted into a gas and used
for fuel.
Black Carbon
A solid form of mostly pure carbon that absorbs solar radiation (light) at all wavelengths. Black
carbon is the most effective form of particulate matter, by mass, at absorbing solar energy, and is
produced by incomplete combustion.
Black Smoke
The term used since the 1950s to describe carbon-containing particulate matter resulting from
incomplete combustion (e.g., from coal); also refers to a measurement that quantified the
concentration of ambient particulate matter. Term has been used as a synonym for soot.
Bottom-Up Inventory
Emissions inventory based on emissions as measured or computed directly by concentration, mass
flow, or stream velocity observations at the source, or calculated (using specific emission factors and
activity levels) on a source-by-source basis.
Boundary Layer
The bottom layer of the atmosphere that is directly influenced by contact with the surface of the
Earth.
Brown Carbon
A class of particulate organic carbon compounds that absorb ultraviolet and visible solar radiation.
BrC can be directly emitted as a product of incomplete combustion, or it can be formed in the
atmosphere as pollutants age.
Carbon Dioxide
A naturally occurring gas which arises from a variety of human activities such as burning fossil
fuels and biomass, land-use changes and other industrial processes. Carbon dioxide is the principal
anthropogenic greenhouse gas that affects the Earth's radiative balance. It is used as the reference
gas against which other greenhouse gases are measured. See also climate change, global warming,
and global warming potential.
Carbon Dioxide Equivalent
A metric used to compare the emissions from various greenhouse gases based upon their global
warming potential. It is the calculated equivalent amount of carbon dioxide emissions that would
result in the same radiative effect as a pulse of emissions of another greenhouse gas. Carbon
dioxide equivalents are commonly expressed as "million metric tons of carbon dioxide equivalents"
(MMTCO2eq).
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Glossary
Carbon Mass Ratio
The ratio of the mass of different components of carbonaceous particles (e.g., the ratio of organic
carbon to black carbon, or the ratio of black carbon to total particulate matter).
Carbonaceous Particulate Matter
A general term for carbon-based compounds found in particles, including black carbon and organic
carbon. Primary combustion particles are largely composed of carbonaceous particulate matter.
Climate Change
Climate change refers to any significant change in measures of climate (such as temperature,
precipitation, or wind) lasting for an extended period (decades or longer). Climate change may
result from natural factors, such as changes in the Sun's intensity or slow changes in the Earth's orbit
around the Sun; natural processes within the climate system (e.g., changes in ocean circulation); and
human activities that change the atmosphere's composition (e.g., through burning fossil fuels) and
the land surface (e.g., deforestation, reforestation, urbanization, desertification, etc.).
Cloud Albedo Effect
The process by which aerosols increase the reflectivity of clouds, leading to negative radiative
forcing.
Cloud Brightening
See cloud albedo effect.
Cloud Burn-off
A subset of the semi-direct effect, related specifically to absorbing aerosols being embedded within
a cloud and resulting in decreased cloud cover.
Cloud Lifetime Effect
The process by which aerosols reduce the size of cloud droplets, resulting in changes to
precipitation patterns and increases in cloud lifetime that lead to cooling.
Coating
Atmospheric process by which a particle can become encased by a shell of a different substance;
this often alters the particle's light absorption qualities.
Co-Emitted Pollutants
Gases and particles that are emitted with black carbon, such as organic carbon, sulfates, nitrates,
sulfur dioxide and nitrogen oxides.
Co-Pollutants
See co-emitted pollutants.
Condensable Particulate Matter
Particles formed as an emissions plume cools in the atmosphere.
Contained Combustion
Closed combustion including internal combustion, reciprocating diesel engines, closed burning.
Differentiated from open burning and open combustion.
Deposition
The transfer of atmospheric gases and particles to the Earth's surface. Wet deposition refers
to deposition that occurs as a result of precipitation. Dry deposition occurs in the absence of
precipitation.
Diesel Particulate Filter
Exhaust emissions control device used to reduce diesel particulate matter; also called a diesel
particulate trap. The diesel particulate filter consists of a porous honeycomb structure that
physically captures and oxidizes the diesel particulate matter.
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Glossary
Diesel Particulate Matter
The particulate component of diesel exhaust, which includes a mixture of black carbon, organic
carbon, sulfates, metals, and trace elements. Black carbon is a major constituent of diesel particulate
matter.
Diesel Retrofit
Any technology or system that achieves emissions reductions beyond those required by new engine
regulations, including the replacement of high-emitting vehicles/equipment with cleaner vehicles/
equipment, repowering or engine replacement, rebuilding the engine to a cleaner standard,
installation of advanced emissions control after-treatment technologies such as DPFs, or the use of a
cleaner fuel.
Direct Effect
The direct scattering or absorption of solar and terrestrial radiation by atmospheric particles.
Direct Forcing
The change in incoming and outgoing solar and terrestrial radiation due to the direct effect of
atmospheric pollutants.
Disability Adjusted Life Years
A measure of overall disease burden, expressed as the number of life years lost due to ill-health,
disability or early death.
Elemental Carbon
A descriptive term for carbonaceous particles that is based on chemical composition rather than
light-absorbing characteristics. Often used as a synonym for black carbon.
Externally Mixed
When each particle in a collection of atmospheric particles is assumed to be composed of only one
chemical compound for purposes of modeling and study.
Fabric Filters
Fabric filters are one of the most widely used devices for controlling emissions of particulate matter.
A fabric filter system typically consists of multiple filter elements (bags) enclosed in a compartment
(housing). When the process stream enters the housing and passes through the filter elements,
particulate matter accumulates as a dust cake on the surface of the bag.
Filter-Based Techniques
One of several ways to quantify the amount of particulate matter in ambient (outdoor) air. Filter-
based measurement methods use samplers that consist of a vacuum pump calibrated to draw in a
fixed volume of air per minute through a filter that captures particles. The average concentration of
particulate matter in the air can be calculated by weighing the filter before and after the run, and
correlating the particulate weight to the volume of air drawn through the pump.
First Indirect Effect
See cloud albedo effect.
Flaming
The stage of combustion when fuel gases are rapidly oxidized. Under oxygen-limited and relatively
low-temperature conditions, soot is emitted.
Fossil Fuels
Fuels derived from coal, oil, and natural gas.
Gasoline Direct Injection
A fuel injection system for gasoline vehicles which introduces fuel directly into each cylinder, which
results in improved fuel economy with higher engine compression rations. Gasoline direct injection
is projected to be used on almost all new model year vehicles starting in 2016.
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Glossary
Glaciation Indirect Effect
A warming effect which occurs in certain mixed-phase clouds when black carbon aerosols (and some
other particles such as mineral dust) serve as ice nuclei in a super-cooled liquid water cloud, thereby
enabling precipitation rather than delaying it.
Global Cost Potential
A metric that compares the relative marginal abatement costs for two climate forcers when a given
climate change target is achieved at least cost.
Global Damage Potential
A metric that compares the relative damage resulting from an equal mass of emissions of two
climate forcers.
Global Temperature Potential
A physical metric that compares the global average temperature change at a given point in time
resulting from equal mass of emissions of two climate forcers.
Global Warming Potential
An index, based upon radiative properties of well-mixed greenhouse gases, measuring the radiative
forcing of a unit mass of a given pollutant in the present-day atmosphere integrated over a chosen
time horizon (often 100 years, or GWP100), relative to that of carbon dioxide (CO2 always has a
global warming potential of 1). The global warming potential represents the combined effect of
the differing times these pollutants remain in the atmosphere and their relative effectiveness in
absorbing radiation.
Greenhouse Gas
Any gas that absorbs infrared radiation in the atmosphere. Greenhouse gases include, but
are not limited to, water vapor, carbon dioxide, methane, nitrous oxide, chlorofluorocarbons,
hydrochlorofluorocarbons, ozone, hydrofluorocarbons, perfluorocarbons, and sulfur hexafluoride.
Hazardous Air Pollutant
Pollutants that are known or suspected to cause cancer or other serious health effects, such as
reproductive effects or birth defects, or adverse environmental effects.
Incomplete Combustion
Combustion where only a partial burning of a fuel occurs. Combustion in practice is almost
always incomplete due to insufficient oxygen or low temperature during the combustion process
preventing the complete oxidation of the fuel to CO2.
Indirect Effects
The various types of absorption or scattering of solar or terrestrial radiation that occur as a result of
anthropogenic aerosol interaction with clouds. These include changes in cloud lifetime, reflectivity,
and composition.
Indirect Forcing
The change in incoming and outgoing solar and terrestrial radiation due to the various indirect
effects resulting from impacts on clouds, including changes in cloud lifetime, reflectivity, and
composition. This forcing can be either positive (warming) or negative (cooling), depending on the
specific cloud interaction.
Infrared Radiation
Radiation emitted by the Earth's surface, the atmosphere, and the clouds. It is also known as
terrestrial or long-wave radiation.
Instantaneous Radiative Forcing
The difference between the amount of radiation coming into the Earth's system and leaving the
system, as measured at the tropopause at a specific instant, due to a change in atmospheric
concentrations. Unlike other radiative forcing measures, instantaneous radiative forcing calculations
do not allow any part of the system to adjust prior to estimating net forcing.
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Glossary
Intergovernmental Panel on Climate Change
Established in 1988 by the World Meteorological Organization and the United Nations Environment
Programme, the Intergovernmental Panel on Climate Change is responsible for providing the
scientific and technical foundation for the United Nations Framework Convention on Climate
Change, primarily through the publication of periodic assessment reports.
Internally Mixed
When individual atmospheric particles are treated as mixtures of chemical components for purposes
of modeling and study.
Kyoto Basket
The set of greenhouse gases covered under the Kyoto Protocol: carbon dioxide, methane, nitrous
oxide, sulfur hexafluoride, hydrofluorocarbons, and perfluorocarbons.
Laser Induced Incandescence
A technique in which a high-energy laser is used to heat soot particles to high temperatures;
measurement of the resulting incandescent light emitted by the soot particles indicates the amount
of soot (soot volume fraction) and its location within the combustion event.
Light-Absorbing Carbon
Carbonaceous particles that absorb light, including black carbon plus brown carbon.
Light-Absorbing Particulate Matter
Refers to particles that tend to absorb light, which represents energy added to the Earth's system
and leads to climate warming.
Light-Scattering Particulate Matter
Refers to particles that tend to reflect or scatter light, which generally leads to increased reflection
of light back to space, causing climate cooling.
Long-Lived Climate Forcer
A pollutant like CO2 that has a positive radiative forcing effect on climate and a long atmospheric
lifetime (decades to centuries).
Maximum Achievable Control Technology (MACT) Standards
U.S. federal emissions standards for stationary sources of hazardous air pollutants requiring the
maximum emissions reductions, taking cost and feasibility into account. Under the Clean Air Act
Amendments of 1990, the MACT must not be less than the average emission level achieved by
controls on the best performing 12 percent of existing sources, by category of industrial and utility
sources. See also National Emissions Standards for Hazardous Air Pollutants (NESHAP).
Metric
An analytical measurement intended to quantify the state of a system. In climate assessments,
metrics are used to quantify the impact of a pollutant relative to a common baseline.
Mixed-Phase Clouds
Clouds with both ice and water.
Mixing
The movement of pollutants through the atmosphere that often leads to chemical and physical
transformations of the pollutants.
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Glossary
National Ambient Air Quality Standards (NAAQS)
Standards required under the Clean Air Act for widespread pollutants from numerous and diverse
sources considered harmful to public health and the environment. Primary standards are designed
to protect public health, including the health of "sensitive" populations such as asthmatics, children,
and the elderly. Secondary standards are designed to protect public welfare, including protection
against decreased visibility, damage to animals, crops, vegetation, and buildings. EPA has set NAAQS
for six principal pollutants, which are called "criteria" pollutants: particulate matter, ozone, lead,
oxides of nitrogen, oxides of sulfur, and carbon monoxide.
National Emissions Inventory
EPA's national emissions database containing information about sources that emit criteria air
pollutants and their precursors, and hazardous air pollutants. The database includes estimates of
annual air pollutant emissions from point, nonpoint, and mobile sources in the 50 states, the District
of Columbia, and Puerto Rico.
National Emissions Standards for Hazardous Air Pollutants (NESHAP)
Stationary source standards for hazardous air pollutants. The NESHAPs promulgated after the 1990
Clean Air Act Amendments require application of technology-based emissions standards referred
to as Maximum Achievable Control Technology (MACT) standards. Consequently, these post-1990
NESHAPs are also referred to as MACT standards.
Net Radiative Forcing
The total radiative forcing due to the presence of a pollutant in the atmosphere, accounting for both
the positive (warming) and negative (cooling) forcing associated with different radiative effects of
the pollutant. For particles, this includes accounting for direct, indirect (cloud), and snow/ice albedo
effects.
New Source Performance Standards (NSPS)
U.S. federal emissions standards for certain air pollutants that are emitted from new, modified,
or reconstructed stationary emissions sources which reflect the use of best available control
technology.
Nitrogen Oxides
A generic term for a group of highly reactive gases, including nitrogen oxide (NO) and nitrogen
dioxide (NO2). Nitrogen oxides result from combustion of fossil or biofuels, especially at high
temperatures.
Open Biomass Burning
Open burning of vegetative material; includes agricultural burning, prescribed burning, and
wildfires.
Organic Carbon
The mix of compounds containing carbon bound with other elements; e.g., hydrogen and oxygen.
Organic carbon may be a product of incomplete combustion, or formed through the oxidation of
VOCs in the atmosphere. Both primary and secondary organic carbon possess radiative properties
that fall along a continuum from light-absorbing to light-scattering.
Organic Carbon to Black Carbon Ratio
See carbon mass ratios.
Organic Matter
The total mass of organic material in a compound.
Oxidation
The chemical reaction of a substance with oxygen or a reaction in which the atoms in an element
lose electrons and its valence is correspondingly increased.
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Glossary
Particle Coagulation
The process by which particles collide and stick together. Part of the internal mixing process,
coagulation reduces the overall particle number and reduces the differences in chemical
composition among the individual particles in the emissions plume.
Particulate Matter
A complex mixture of extremely small particles and liquid droplets suspended in the atmosphere.
Particulate matter (PM) is made up of a number of components, including acids (such as nitrates and
sulfates), organic chemicals, metals, and soil or dust particles. For purposes of air quality and health
studies, PM is typically measured in two size ranges: PMi0 and PM2.5.
PM10
Particulate matter with an aerodynamic diameter less than or equal to a nominal 10 micrometers.
PMio includes PM2.5.
PM2.5
Fine particulate matter with an aerodynamic diameter less than or equal to 2.5 micrometers.
Photoacoustic
A black carbon measurement technique where light from a source is absorbed by the aerosol
resulting in the heating and subsequent expansion of the surrounding air. The expansion results in a
sound wave which is then detected with a microphone.
Polycyclic Aromatic Hydrocarbon
A group of organic contaminants formed from incomplete combustion. These aromatic compounds
comprise two or more benzene rings arranged in various configurations (polycyclic). Many
polycyclic aromatic hydrocarbons (PAHs) are known to be toxic to humans and ecosystems, and
EPA has classified seven PAH compounds as probable human carcinogens: benzo[a]pyrene, benz[a]
anthracene, chrysene, benzo[b]fluoranthene, benzo[k]fluoranthene, dibenz[a,h]anthracene, and
indeno[l,2,3cd]pyrene). PAHs are precursors (building blocks) in the formation of black carbon and
brown carbon.
Prescribed Fire
Any fire intentionally ignited by management under an approved plan to meet specific objectives.
Primary Particle
A particle that is emitted directly from a source.
Pulse
See specific forcing pulse.
Pyrolysis
The heating of solid fuels in the absence of oxygen. Pyrolysis induces the evaporation of volatile
gases from the solid fuel needed to support combustion. Thermal breakdown of portions of the
solid fuel provide additional fuel gases. Pyrolysis is used to produce charcoal and biochar, a residual
form of carbon in solid form.
Radiation
Energy in the form of electromagnetic waves (or photons). Photons absorbed by solid materials,
such as light-absorbing particles, are transformed into other forms of energy. Most notably, solar
radiation absorbed by a particle converts, in part, into heat energy that warms the surrounding
atmosphere or the surfaces upon which the particles are deposited (e.g., snow and ice).
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Glossary
Radiative Forcing
The change in the energy balance between incoming solar radiation and exiting infrared radiation,
typically measured in watts per square meter (W m-2), due to a change in concentration (generally
the change since preindustrial conditions in 1750). Positive radiative forcing tends to warm the
surface of the Earth, while negative forcing generally leads to cooling. A pollutant that increases the
amount of energy in the Earth's climate system is said to exert "positive radiative forcing," which
leads to warming. In contrast, a pollutant that exerts "negative radiative forcing" reduces the amount
of energy in the Earth's system and leads to cooling.
Second Indirect Effect
See cloud lifetime effect.
Secondary Organic Aerosols
Carbonaceous aerosols that are produced in the atmosphere rather than being directly emitted.
Precursor gases (such as aromatic hydrocarbons, monoterpenes) undergo chemical reactions and
condensation to form secondary organic aerosols.
Secondary Particle
A particle (e.g., sulfate or nitrate) that is formed in the atmosphere from the oxidation of gaseous
precursors like sulfur dioxide, nitrogen oxides, and volatile organic compounds or through the
transformation of directly emitted particles. The acids resulting from the oxidation of these
compounds attract water vapor to form tiny droplets (fine particles).
Semi-Direct Effect
Localized heating of the atmosphere by absorbing aerosol particles, affecting the relative humidity
and stability of the troposphere, which in turn affect cloud formation and lifetime.
Short-Lived Climate Forcer
A pollutant, such as black carbon, ozone, or methane, that has a positive radiative forcing effect on
climate but a relatively short atmospheric lifetime (days to years).
Single-Scattering Albedo
The ratio of scattering optical depth to the total optical depth (scattering plus absorption) of the
atmosphere; indicates how much of the light extinction in the atmosphere is due to scattering vs.
absorption. If single-scattering albedo equals 1, all particle extinction is due to scattering; if single-
scattering albedo equals 0, all extinction is due to absorption.
Smoldering
A non-flaming phase of the combustion process that involves a slower, cooler form of combustion
which occurs as oxygen attacks the surface of heated solid fuel directly.
Snow/Ice Albedo Effect
Decrease in reflectivity (and increase in absorption) of solar radiation that occurs as a result of the
darkening of snow and ice through aerosol deposition.
Snow/Ice Albedo Forcing
Positive radiative forcing resulting from the deposition of black carbon on snow and ice, which
darkens the surface and decreases reflectivity (albedo), thereby increasing absorption of solar
radiation and accelerating melting.
Social Cost of Carbon
An estimate of the monetized damages resulting from an incremental increase in CO2 emissions in a
given year. It can be thought of as the monetized benefit to society of reducing one ton of CO2.
xxviii Report to Congress on Black Carbon
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Glossary
Solar Radiation
Radiation emitted by the Sun. It is also referred to as short-wave radiation. Solar radiation has a
distinctive range of wavelengths (spectrum) determined by the temperature of the Sun. In the
context of this report, the term refers to the portions of the solar spectrum which reach the lower
atmosphere, including the ultraviolet (>280 nm), visible and infrared.
Solar Zenith Angle
The angle between a point directly above any location on the Earth's surface (the zenith) and the
Sun, as measured at the location. These angles relate to the elevation of the Sun above the horizon
(in degrees).
Soot
A complex mixture of mostly black and organic carbon that is the primary light-absorbing pollutant
emitted by the incomplete combustion of fossil fuels, biofuels, and biomass.
Source Apportionment
The use of ambient and/or emissions data along with statistical modeling to determine the
contribution of a specific emissions source category to measured ambient concentrations of air
pollutants like PM2.5.
SPECIATE database
EPA's repository of speciation profiles characterizing the composition of emissions from specific air
pollution sources.
Specific Forcing Pulse
A metric based on the amount of energy added to the Earth's system due to the radiative forcing
caused by a given mass of a pollutant.
Surface Dimming Effect
The reduction of solar radiation at Earth's surface due to high concentrations of particles, especially
light absorbing particles, in the atmosphere, above. This results in cooling at the Earth's surface
(even though net forcing measured at the top of the atmosphere might be positive).
Surface Temperature Response per Unit Continuous Emission
A metric that compares the change in surface temperature due to an assumed continuous emission
of equal masses a climate forcer and CO2.
Tar Balls
Liquid aerosol droplets observed in biomass burning plumes that appear to be formed entirely from
brown carbon.
Thermal Techniques
A variety of approaches used for the measurement of organic and elemental carbon. The techniques
involve the heating of particulate matter with the subsequent detection of the evolved carbon with a
variety of techniques.
Thermal-Optical Techniques
A variety of approaches used for the measurement of organic and elemental carbon. These
techniques are similar to the thermal techniques defined above, with the addition of an optical
measurement to improve the separation of elemental carbon from organic carbon.
Thermodynamic Effect
The process by which freezing of droplets in mixed-phase clouds is delayed because droplet size
is reduced. This effect ultimately changes the characteristics of the cloud, but whether it leads to
warming or cooling is unclear.
Third Pole
Refers to the Hindu Kush-Himalayan-Tibetan region.
Report to Congress on Black Carbon xxix
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Glossary
Tier 2 Standards
U.S. EPA emissions standards for emissions of hydrocarbons, carbon monoxide, oxides of nitrogen,
and PM from light-duty vehicles (automobiles and light-duty trucks) phased in with the 2004
through 2006 model years. Standards represent roughly 99% emissions reduction compared to pre-
control (pre-1968 model year) vehicles.
Top-Down Inventory
Emissions inventory based on ambient air quality data. Individual emissions source estimates are
based upon the relative magnitude of tracer compounds in the ambient air.
Top of the Atmosphere
The location between the troposphere and the stratosphere. Measuring radiative forcing at this
altitude is best for determining net energy balance.
Top of the Atmosphere Radiative Forcing
Net radiative forcing measured (or modeled) at the top of the atmosphere to capture the total
change in incoming and outgoing radiation due to the presence of atmospheric pollutants.
Transport
See atmospheric transport.
Troposphere
The lowest part of the atmosphere from the Earth's surface (ranging from 9 km in high latitudes to
16 km in the tropics on average) where clouds and "weather" phenomena occur. In the troposphere,
temperatures generally decrease with height.
Ultra-Low Sulfur Diesel Fuel
Diesel fuel that has a maximum of 15 parts per million of sulfur content. Ultra-low sulfur diesel
enables advanced pollution control technologies such as diesel paniculate filters and urea selective
catalytic reduction systems for NOx.
Ultraviolet Radiation
The energy range just beyond the violet end of the visible spectrum, with wavelengths between 10-
400 nm. Ultraviolet radiation constitutes only about 5 percent of the total energy emitted from the
Sun, and most ultraviolet radiation is blocked by Earth's atmosphere. The ultraviolet radiation that
does reach the Earth aids in plant photosynthesis and helps produce vitamin D in humans. Too much
ultraviolet radiation can burn the skin, cause skin cancer and cataracts, and damage vegetation.
Value of a Statistical Life
A summary measure for the dollar value of small changes in mortality risk experienced by a large
number of people. The total estimated value of a statistical life is derived from aggregated estimates
of individual values for small changes in mortality risks.
Volatile Organic Compounds
Organic chemical compounds whose composition makes it possible for them to evaporate under
normal atmospheric conditions of temperature and pressure. These carbonaceous pollutant gases
are emitted by both anthropogenic and natural processes and often serve as precursors for the
formation of aerosol particles and ozone. Examples of volatile organic compounds include benzene,
toluene, methylene chloride, and methyl chloroform.
Wildfire
An unplanned ignition caused by lightning, volcanoes, unauthorized activity, accidental human
actions, and escaped prescribed fires.
Willingness to Pay
The maximum amount a person would be willing to pay, sacrifice, or exchange in order to receive a
good or to avoid something undesired, such as pollution.
xxx Report to Congress on Black Carbon
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Executive Summary
Black carbon (BC) emissions have important
impacts on public health, the environment, and
the Earth's climate. BC is a significant component of
particle pollution, which has been linked to adverse
health and environmental impacts through decades
of scientific research. Recent work indicates that
BC also plays an important role in climate change,
although there is more uncertainty about its effects
on climate than for greenhouse gases (GHG), such
as carbon dioxide and methane. BC has been linked
to a range of climate impacts, including increased
temperatures, accelerated ice and snow melt, and
disruptions to precipitation patterns. Importantly,
reducing current emissions of BC may help slow
the near-term rate of climate change, particularly
in sensitive regions such as the Arctic. However,
BC reductions cannot substitute for reductions in
long-lived GHGs, which are necessary for mitigating
climate change in the long run.
Despite the rapidly expanding body of scientific
literature on BC, there is a need for a more
comprehensive evaluation of both the magnitude
of particular global and regional climate effects due
to BC and the impact of emissions mixtures from
different source categories. To advance efforts to
understand the role of BC in climate change, on
October 29, 2009, Congress requested the U.S.
Environmental Protection Agency (EPA) conduct a BC
study as part of /-/./?. 2996: Department of the Interior,
Environment, and Related Agencies Appropriations
Act, 2010. Specifically, the legislation stated that:
"Not later than 18 months after the date of
enactment of this Act, the Administrator, in
consultation with other Federal agencies, shall
carry out and submit to Congress the results of a
study on domestic and international black carbon
emissions that shall include
• an inventory of the major sources of black carbon,
• an assessment of the impacts of black carbon on
global and regional climate,
• an assessment of potential metrics and
approaches for quantifying the climatic effects
of black carbon emissions (including its radiative
forcing and warming effects) and comparing
those effects to the effects of carbon dioxide and
other greenhouse gases,
• an identification of the most cost-effective
approaches to reduce black carbon emissions,
and
• an analysis of the climatic effects and other
environmental and public health benefits of
those approaches."
To fulfill this charge, EPA has conducted an intensive
effort to compile, assess, and summarize available
scientific information on the current and future
impacts of BC, and to evaluate the effectiveness of
available BC mitigation approaches and technologies
for protecting climate, public health, and the
environment. As requested by Congress, EPA
has consulted with other federal agencies on key
elements of this report, including inventories, health
and climate science, and mitigation options. The
report draws from recent BC assessments, including
work under the United Nations Environment
Programme (UNEP) and the World Meteorological
Organization (WMO), the Convention on Long
Range Transboundary Air Pollution (CLRTAP), and
the Arctic Council. Each of these individual efforts
provides important information about particular
sectors, regions, or issues. The task outlined for EPA
by Congress is broader and more encompassing,
requiring a synthesis of currently available
information about BC across numerous bodies
of scientific inquiry. The results are presented in
this Report to Congress on Black Carbon. The key
messages of this report can be summarized as
follows.
1. Black carbon is the most strongly light-
absorbing component of particulate matter
(PM), and is formed by the incomplete
combustion of fossil fuels, biofuels, and
biomass.
BC can be defined specifically as a solid form of
mostly pure carbon that absorbs solar radiation
(light) at all wavelengths. BC is the most effective
form of PM, by mass, at absorbing solar energy;
Report to Congress on Black Carbon
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Executive Summary
Global BC Emissions, 2000 (7,600 Gg)
0.5% 0.7%
19.3%
U.S. BC Emissions in 2005 (0.64 Million Tons)
1.1%
35.5%
52.3%
19.0%
35.3%
25.1%
1.0%
6.8%
• Open Biomass Burning Domestic/Residential
1 (Includes Wildfires)
Transp°rt Other
• Energy/Power
Figure A. BC Emissions by Major Source Category. (Source: Lamarque et al., 2010 and U.S. EPA)
other types of particles, including sulfates, nitrates
and organic carbon (OC), generally reflect light. BC
is a major component of "soot," a complex light-
absorbing mixture that also contains organic carbon.
Recent estimates of BC emissions by source category
in the United States and globally are shown in Figure
A.
2. BC is emitted directly into the atmosphere in
the form of fine particles (PM2.S). The United
States contributes about 8% of the global
emissions of BC. Within the United States,
BC is estimated to account for approximately
12% of all direct PM2_5 emissions in 2005.
Many countries have significantly higher
PM2.S emissions than the United States,
and countries with a different portfolio of
emissions sources might have a significantly
higher percentage of BC.
3. BC contributes to the adverse impacts on
human health, ecosystems, and visibility
associated with PM2.S.
Short-term and long-term exposures to PM2.5 are
associated with a broad range of human health
impacts, including respiratory and cardiovascular
effects, as well as premature death. PM2.5, both
ambient and indoor, is estimated to result in millions
of premature deaths worldwide, the majority of
which occur in developing countries. The World
Health Organization estimates that indoor smoke
from solid fuels is the 10th major mortality risk
factor globally, contributing to approximately 2
million deaths annually. Women and children are
particularly at risk. Ambient air pollution is also a
significant health threat: according to the WHO,
urban air pollution is among the top ten risk factors
in medium- and high-income countries. Urban air
pollution is not ranked in the top ten major risk
factors in low-income countries since other risk
factors (e.g., childhood underweight and unsafe
water, sanitation and hygiene) are so substantial;
however, a much larger portion of the total deaths
related to ambient PM2.5 globally are expected to
occur in developing regions, partly due to the size of
exposed populations in those regions. PM2.5 is also
linked to adverse impacts on ecosystems, to visibility
impairment, to reduced agricultural production in
some parts of the world, and to materials soiling and
damage.
Over the past decade, the scientific community
has focused increasingly on trying to identify the
health impacts of particular PM2.5 constituents, such
as BC. However, EPA has determined that there is
insufficient information at present to differentiate
the health effects of the various constituents of
PM25; thus, EPA assumes that many constituents
are associated with adverse health impacts. It
is noteworthy that emissions and ambient
concentrations of directly emitted PM2.5 are often
highest in urban areas, where large numbers of
people live.
Report to Congress on Black Carbon
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Executive Summary
4. BC influences climate through multiple
mechanisms:
• Direct effect: BC absorbs both incoming and
outgoing radiation of all wavelengths, which
contributes to warming of the atmosphere and
dimming at the surface.
• Snow/ice albedo effect: BC deposited on snow and
ice darkens the surface and decreases reflectivity,
thereby increasing absorption and accelerating
melting.
• Other effects: BC also alters the properties of
clouds, affecting cloud reflectivity and lifetime
("indirect effects"), stability ("semi-direct effect")
and precipitation.
5. The direct and snow/ice albedo effects of
BC are widely understood to lead to climate
warming. However, the globally averaged net
climate effect of BC also includes the effects
associated with cloud interactions, which
are not well quantified and may cause either
warming or cooling. Therefore, though most
estimates indicate that BC has a net warming
influence, a net cooling effect cannot be ruled
out. It is also important to note that the net
radiative effect of all aerosols combined
(including sulfates, nitrates, BC and OC) is
widely understood to be negative (cooling) on
a global average basis.
The direct radiative forcing effect of BC is the best
quantified and appears to be positive and significant
Black carbon direct TOA forcing (W nr2)
90,= , , , =, —5
45
0
-45
-90
90
45
-45
-90
Black carbon cryosphere forcing (W rrr2)
5
2
1
0.5
0.25
0.1
0.05
0.025
Figure B. Regional Variability in Direct Radiative Forcing and Snow/Ice Albedo Forcing for BC from All
Sources, simulated with the Community Atmosphere Model. (Source: Bond et al., 2011)
Report to Congress on Black Carbon
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Executive Summary
on both global and regional scales. This warming
effect is augmented by deposition of BC on snow and
ice. These effects are shown in Figure B. The central
estimates of global average direct forcing by BC from
surveyed studies range from +0.34 to +1.0 Watts
per square meter (W rrr2). A recent UNEP/WMO
assessment presented a narrower central range of
+0.3 to +0.6 W m2. These estimates are generally
higher than the 2007 Intergovernmental Panel on
Climate Change (IPCC) estimate of +0.34 (±0.25) W
m2.
The snow/ice albedo effect from BC has been
estimated in recent studies to add about +0.05
W m-2, generally less than the +0.1 (±0.1) W m-2
estimated by the IPCC; however, UNEP/WMO found
that when the snow/ice albedo forcing estimates are
adjusted to account for the greater warming efficacy
of the snow/ice deposition mechanism, the snow/
ice albedo effect could add +0.05 to +0.25 W m2 of
forcing. The sum of the direct and snow/ice albedo
effects of BC on the global scale is likely comparable
to or larger than the forcing effect from methane,
but less than the effect of carbon dioxide;1 however,
there is more uncertainty in the forcing estimates for
BC.
The climate effects of BC via interactions with clouds
are more uncertain, and their net climate influence
is not yet clear. All aerosols (including BC) affect
climate indirectly by changing the reflectivity
(albedo) and lifetime of clouds. The net indirect
effect of all aerosols is very uncertain but is thought
to have a net cooling influence. The IPCC estimated
the global average cloud albedo forcing from all
aerosols as -0.7 W m2 (with a 5 to 95% confidence
range of -0.3 W m2 to -1.80 W rrr2). The IPCC did
not provide quantitative estimates of the effect of
aerosols on cloud lifetime, and the contribution of
BC to these indirect effects has not been explicitly
quantified to date. BC has additional effects on
clouds—including changes to cloud stability and
enhanced precipitation from colder clouds—that
can lead to either warming or cooling. However, few
quantitative estimates of these effects are available,
and significant uncertainty remains. Due to all of
the remaining gaps in scientific knowledge, it is
difficult to place quantitative bounds on the forcing
attributable to BC impacts on clouds at present;
however, UNEP/WMO have provided a central
forcing estimate of -0.4 to +0.4 W rrr2 for all of the
cloud effects of BC combined.
The sign and magnitude of the net climate forcing
from BC emissions are not fully known at present,
1 The IPCC's radiative forcing estimates for elevated concentrations
of CO2 and methane are +1.66 W m 2 and +0.48 W m 2, respectively.
largely due to remaining uncertainties regarding the
effects of BC on clouds. There is inconsistency among
reported observational and modeling results,
and many studies do not provide quantitative
estimates of cloud impacts. In the absence of a full
quantitative assessment, the current scientific basis
for understanding BC climate effects is incomplete.
Based on a limited number of modeling studies,
the recent UNEP/WMO assessment estimated that
global average net BC forcing is likely to be positive
and in the range of 0.0 to +1.0 W rrr2, with a best
estimate of +0.6 W rrr2; however, further work is
needed to refine these estimates.
6. Sensitive regions such as the Arctic and the
Himalayas are particularly vulnerable to the
warming and melting effects of BC.
Studies have shown that BC has especially strong
impacts in the Arctic, contributing to earlier spring
melting and sea ice decline. All particle mixtures
reaching the Arctic are a concern, because even
emissions mixtures that contain more reflective
(cooling) aerosols can lead to warming if they are
darker than the underlying ice or snow. Studies
indicate that the effect of BC on seasonal snow
cover duration in some regions can be substantial,
and that BC deposited on ice and snow will continue
to have radiative effects as long as the BC remains
exposed (until the snow melts away or fresh snow
falls). BC has also been shown to be a significant
factor in the observed increase in melting rates of
some glaciers and snowpack in parts of the Hindu
Kush-Himalayan-Tibetan (HKHT) region (the "third
pole").
7. BC contributes to surface dimming, the
formation of Atmospheric Brown Clouds
(ABCs), and changes in the pattern and
intensity of precipitation.
The absorption and scattering of incoming solar
radiation by BC and other particles cause surface
dimming by reducing the amount of solar radiation
reaching the Earth's surface. In some regions,
especially Asia, southern Africa, and the Amazon
Basin, BC, sulfates, organics, dust and other
components combine to form pollution clouds
known as Atmospheric Brown Clouds (ABCs). ABCs
have been linked to surface dimming and a decrease
in vertical mixing, which exacerbates air pollution
episodes. ABCs also contribute to changes in the
pattern and intensity of rainfall, and to observed
changes in monsoon circulation in South Asia. In
general, regional changes in precipitation due to BC
and other aerosols are likely to be highly variable,
with some regions seeing increases while others
experience decreases.
Report to Congress on Black Carbon
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Executive Summary
BC, 2000
0.1 0.2 0.5 1 2 5 10 20
Figure C. BC Emissions, 2000, Gg. (Courtesy of Tami Bond, produced based on data from Bond et al., 2007)
8. BC is emitted with other particles and gases,
many of which exert a cooling influence
on climate. Therefore, estimates of the net
effect of BC emissions sources on climate
should include the offsetting effects of these
co-emitted pollutants. This is particularly
important for evaluating mitigation options.
Some combustion sources emit more BC
than others relative to the amount of co-
pollutants; reductions from these sources have
the greatest likelihood of providing climate
benefits.
The same combustion processes that produce BC
also produce other pollutants, such as sulfur dioxide
(SO2), nitrogen oxides (NOX), OC and CO2. Some of
these co-emitted pollutants result in "scattering" or
reflecting particles (e.g. sulfate, nitrate, OC) which
exert a cooling effect on climate. The sign and
magnitude of the forcing resulting from particular
emissions mixtures depend on their composition.
For example, the particles emitted by mobile diesel
engines are about 75% BC, while particle emissions
from biomass burning are dominated by OC. Sources
rich in BC have a greater likelihood of contributing to
climate warming, and this may affect climate-related
mitigation choices. Although OC generally leads to
cooling, some portion of co-emitted OC, notably
brown carbon (BrC), partially absorbs solar radiation.
The net contribution of BrC to climate is presently
unknown.
Atmospheric processes that occur after BC is
emitted, such as mixing, aging, and coating, can also
affect the net influence on climate.
9. BC's short atmospheric lifetime (days to
weeks), combined with its strong warming
potential, means that targeted strategies
to reduce BC emissions can be expected to
provide climate benefits within the next
several decades.
Because the duration of radiative forcing by BC is
very limited, the climate will respond quickly to BC
emissions reductions, and this can help slow the
rate of climate change in the near term. In contrast,
long-lived GHGs may persist in the atmosphere
for centuries. Therefore, reductions in GHG
emissions will take longer to influence atmospheric
concentrations and will have less impact on climate
on a short timescale. However, since GHGs are the
largest contributor to current and future climate
change, and because GHGs accumulate in the
atmosphere, deep reductions in these pollutants are
necessary for limiting climate change over the long-
term.
Emissions sources and ambient concentrations of
BC vary geographically and temporally (Figure C),
resulting in climate effects that are more regional
and seasonal than the more uniform effects of
long-lived, well-mixed GHGs. Likewise, mitigation
actions for BC will produce different climate results
depending on the region, season, and sources in the
area where emissions reductions occur.
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Executive Summary
10. The different climate attributes of BC and
long-lived GHGs make it difficult to interpret
comparisons of their relative climate impacts
based on common metrics.
Due in large part to the difference in lifetime
between BC and CO2, a comparison between the
relative climate impacts of BC and CO2 (or other
climate forcers) is very sensitive to the metric used.
There is currently no single metric (e.g., Global
Warming Potential or GWP) that is widely accepted
by the science and research community for this
purpose. However, new metrics designed specifically
for short-lived climate forcers like BC have recently
been developed, and these metrics may enable
better prioritization among mitigation options with
regard to potential net climate effects.
11. Based on recent emissions inventories (2000
for global and 2005 for the United States), the
majority of global BC emissions come from
Asia, Latin America, and Africa. The United
States currently accounts for approximately
8% of the global total, and this fraction is
declining. Emissions patterns and trends
across regions, countries and sources vary
significantly.
Though there is significant uncertainty in global BC
emissions inventories, recent studies indicate that
global BC emissions have been increasing for many
decades. However, emissions of BC in North America
and Europe have declined substantially since the
early 1900s and are expected to decline further in
the next several decades due to pollution controls
and use of cleaner fuels. Elsewhere, BC emissions
have been increasing, with most of the increase
coming from developing countries in Asia, Africa
and Latin America. According to available estimates,
these regions currently contribute more than 75%
of total global BC emissions, with the majority
of emissions coming from the residential sector
(cookstoves) and open biomass burning. Current
emissions from the United States, OECD Europe,
the Middle East, and Japan come mainly from the
transportation sector, particularly from mobile diesel
engines. In the United States, nearly 50% of BC
emissions came from mobile diesel engines in 2005.
12. Control technologies are available to reduce
BC emissions from a number of source
categories.
BC emissions reductions are generally achieved by
applying technologies and strategies to improve
combustion and/or control direct PM2.5 emissions
from sources. Though the costs of such mitigation
approaches vary, many reductions can be achieved at
reasonable costs. Controls applied to reduce BC will
help reduce total PM2.5 and other co-pollutants.
13. BC mitigation strategies, which lead to
reductions in fine particles, can provide
substantial public health and environmental
benefits.
Strategies to reduce BC generally lead to reductions
in emissions of all particles from a particular source.
Thus, while it is not easy to reduce BC in isolation
from other constituents, most mitigation strategies
will provide substantial benefits in the form of
PM2.5 reductions. Reductions in directly emitted
PM25 can substantially reduce human exposure,
providing large public health benefits that often
exceed the costs of control. In the United States,
the average public health benefits associated with
reducing directly emitted PM2.5 are estimated to
range from $290,000 to $1.2 million per ton PM2.5
in 2030 (2010$). The cost of the controls necessary
to achieve these reductions is generally far lower.
For example, the costs of PM controls for new diesel
engines are estimated to be about $14,000 per ton
PM2.5 (2010$). BC reduction strategies implemented
at the global scale could provide very large benefits:
the PM25 reductions resulting from BC mitigation
measures could potentially result in hundreds of
thousands of avoided premature deaths each year.
14. Mitigating BC can also make a difference
in the short term for climate, at least in
sensitive regions.
Benefits in sensitive regions like the Arctic, or in
regions of high emissions such as Asia, may include
reductions in warming and melting (ice, snow,
glaciers) and reversal of changes in precipitation
patterns. BC reductions could help reduce the
rate of warming soon after they are implemented.
However, available studies also suggest that BC
mitigation alone would be insufficient to change
the long-term trajectory of global warming (which is
driven by GHGs).
15. Selecting optimal BC mitigation measures
requires taking into account the full suite
of impacts and attempting to maximize
co-benefits and minimize unintended
consequences across all objectives (health,
climate, and environment).
With a defined set of goals, policymakers can
evaluate the "mitigation potential" within each
country or region. The mitigation potential depends
on total BC emissions and key emitting sectors,
and also depends on the availability of control
technologies or alternative mitigation strategies.
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Executive Summary
POTENTIAL BENEFITS = MITIGATION POTENTIAL +/- CONSTRAINING FACTORS
Goals
Climate
Radiative Forcing
Temperature
Ice/Snow Melt
Precipitation
Health
Ambient Exposures
Indoor Exposures
Environment
Surface Dimming
Visibility
Emissions sources
Stationary
Sources
Brick Kilns
Coke Ovens
Diesel Generator:
Utilities
Flaring
Mobile
Sources
On-Road Diesel
On-Road Gasoline
Construction Equip.
Agricultural Equip.
Locomotives
Marine
Open Biomass
Burning
Agricultural Burning
Prescribed Burning
Residential
Cooking and
Heating
Cookstoves
Wood stoves
Hydronic Heaters
Mitigation options
Available Control
Technologies
e.g. Diesel
Participate Filters
Alternative Strategies
to Reduce Emissions
e.g. Efficiency
Improvements, Substitution
Timing
Location
Atmospheric
Transport
Co-Emitted
Pollutants
Cost
Existing Regulatory
Programs
Implementation
Barriers
Uncertainty
Figure D. Policy Framework for Black Carbon Mitigation Decisions. (Source: U.S. EPA.)
As illustrated in Figure D, the ideal emissions
reduction strategies will also depend on a range of
constraining factors, including:
• Timing
• Location
• Atmospheric Transport
• Co-emitted Pollutants
• Cost
• Existing Regulatory Programs
• Implementation Barriers
• Uncertainty
16. Considering the location and timing of
emissions and accounting for co-emissions
will improve the likelihood that mitigation
strategies will be properly guided by the
balance of climate and public health
objectives.
PM mitigation strategies that focus on sources
known to emit large amounts of BC—especially
those with a high ratio of BC to OC, like diesel
emissions—will maximize climate co-benefits. The
timing and location of the reductions are also very
important. Some of the most significant climate
benefits of BC-focused control strategies may
come from reducing emissions affecting the Arctic,
Himalayas and other ice and snow-covered regions.
The effect of BC emissions reductions on human
health is a function of changing exposure and the
size of the affected population. The largest health
benefits from BC-focused control strategies will
occur locally near the emissions source and where
exposure affects a large population.
17. Achieving further BC reductions, both
domestically and globally, will require adding
a specific focus on reducing direct PM2.S
emissions to overarching fine particle control
programs.
BC reductions that have occurred to date (largely
in developed countries) are mainly due to control
programs aimed at PM2.5, not targeted efforts to
reduce BC per se. Greater attention to BC-focused
strategies has the potential to help protect the
climate (via the BC reductions achieved through
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Executive Summary
direct PM2.5 controls) while ensuring continued
improvements in public health (via control of direct
PM2.5in highly populated areas). Even if such controls
are more costly than controls on secondary PM
precursors, the combined public health and climate
benefits may justify the expense.
18. The most promising mitigation options
identified in this report for reducing BC (and
related "soot") emissions are consistent with
control opportunities emphasized in other
recent assessments.
• United States: The United States will achieve
substantial BC emissions reductions by 2030,
largely due to controls on new mobile diesel
engines. Diesel retrofit programs for in-use
mobile sources are a valuable complement to new
engine standards for reducing emissions. Other
source categories in the United States, including
stationary sources (industrial, commercial and
institutional boilers, stationary diesel engines,
uncontrolled coal-fired electric generating units),
residential wood combustion (hydronic heaters
and woodstoves), and open biomass burning also
offer potential opportunities but have more limited
mitigation potential due to smaller remaining
emissions in these categories, or limits on the
availability of effective BC control strategies.
- Total mobile source BC emissions are
projected to decline by 86% by 2030 due
to regulations already promulgated. BC
emissions from mobile diesel engines
(including on-road, non road, locomotive, and
commercial marine engines) in the United
States are being controlled through two
primary mechanisms: (1) emissions standards
for new engines, including requirements
resulting in use of diesel particulate filters
(DPFs) in conjunction with ultra low sulfur
diesel fuel; and (2) retrofit programs for
in-use mobile diesel engines, such as EPA's
National Clean Diesel Campaign and the
SmartWay Transport Partnership Program.
Substantial future reductions in mobile diesel
emissions are anticipated through new engine
requirements and diesel retrofit programs.
- BC emissions from stationary sources in the
United States have declined dramatically in
the last century, with remaining emissions
coming primarily from coal combustion
(utilities, industrial/commercial boilers, other
industrial processes) and stationary diesel
engines. Available control technologies and
strategies include use of cleaner fuels and
direct PM2.5 reduction technologies such
as fabric filters (baghouses), electrostatic
precipitators (ESPs), and DPFs.
- Emissions of all pollutants from residential
wood combustion (RWC) are currently
being evaluated as part of EPA's ongoing
review of emissions standards for residential
wood heaters, including hydronic heaters,
woodstoves, and furnaces. Mitigation options
include providing alternatives to wood,
replacing inefficient units or retrofitting
existing units.
- Open biomass burning, including both
prescribed fires and wildfires, represents a
potentially large but less certain portion of
the U.S. BC inventory. These sources emit
much larger amounts of OC compared to
BC. The percent of land area affected by
different types of burning is uncertain, as are
emissions estimates. Appropriate mitigation
measures depend on the timing and location
of burning, resource management objectives,
vegetation type, and available resources. For
wildfires, expanding domestic fire prevention
efforts may help to reduce BC emissions.
Global: The most important BC emissions
reduction opportunities globally include residential
cookstoves in all regions; brick kilns and coke ovens
in Asia; and mobile diesels in all regions. A variety
of other opportunities may exist in individual
countries or regions.
- Other developed countries have emissions
patterns and control programs that are
similar to the United States, though the
timing of planned emissions reductions may
vary. Developing countries have a higher
concentration of emissions in the residential
and industrial sectors, but the growth of
the mobile source sector in these countries
may lead to an increase in their overall
BC emissions and a shift in the relative
importance of specific BC-emitting sources
over the next several decades.
- For mobile sources, both new engine
standards and retrofits of existing engines/
vehicles may help reduce BC emissions in
the future. While many other countries have
already begun phasing in emissions and fuel
standards, BC emissions in this category
in developing countries are expected to
continue to increase. Emissions control
requirements lag behind in some regions,
as does on-the-ground deployment of DPFs
and low sulfur fuels. Further or more rapid
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Executive Summary
reductions in BC will depend on accelerated
deployment of clean engines and fuels.
Emissions from residential cookstoves are
both a large source of BC globally and a
major threat to public health. Approximately
3 billion people worldwide cook their food or
heat their homes by burning biomass or coal
in rudimentary stoves or open fires, resulting
in pollution exposures that lead to 2 million
deaths each year. Mitigation in this sector
represents the area of largest potential public
health benefit of any of the sectors considered
in this report. Significant expansion of current
clean cookstove programs would be necessary
to achieve large-scale climate and health
benefits. A wide range of improved stove
technologies is available, but the potential
climate and health benefits vary substantially
by technology and fuel. Setting BC emissions
reductions as a policy priority would drive
cookstove efforts toward solutions that
achieve this goal. A number of factors point to
much greater potential to achieve large-scale
success in this sector today.
The largest stationary sources of BC
emissions internationally include brick
kilns, coke ovens (largely from iron/
steel production), and industrial boilers.
Replacement or retrofit options already exist
for many of these source categories.
Open biomass burning is the largest source
of BC emissions globally. However, emissions
of OC (including potentially light absorbing
BrC) are approximately seven times higher
than BC emissions from this sector, and
more complete emissions inventory data are
needed to characterize impacts of biomass
burning and evaluate the effectiveness
of mitigation measures at reducing BC.
Expanded wildfire prevention efforts may
help to reduce BC emissions globally.
Successful implementation of mitigation
approaches in world regions where biomass
burning is widespread will require training
in proper burning techniques and tools to
ensure effective use of prescribed fire.
• Sensitive Regions: To address impacts in the
Arctic, other assessments have identified the
transportation sector (land-based diesel engines
and Arctic shipping); residential heating (wood-
fired stoves and boilers); and forest, grassland
and agricultural burning as primary mitigation
opportunities. In the Himalayas, studies have
focused on residential cooking; industrial
sources (especially coal-fired brick kilns); and
transportation, primarily on-road and off-road
diesel engines.
19. A variety of other options may also be
suitable and cost-effective for reducing BC
emissions, but these can only be identified
with a tailored assessment that accounts for
individual countries' resources and needs.
Some potential sectors of interest for further
exploration include agricultural burning, oil and gas
flaring, and stationary diesel engines in the Arctic far
north.
20. Despite some remaining uncertainties about
BC that require further research, currently
available scientific and technical information
provides a strong foundation for making
mitigation decisions to achieve lasting
benefits for public health, the environment,
and climate.
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Chapter 1
Introduction
Black carbon (BC) has recently received a
great deal of attention among scientists and
policymakers for its impacts on global and regional
climate. Though substantial and immediate
reductions in long-lived greenhouse gases (GHG)
are essential for solving the problem of climate
change over the long term, BC offers a promising
mitigation opportunity to address climate effects in
the near-term and to slow the rate of climate change.
BC's high capacity for light absorption and its role
in key atmospheric processes link it to a range of
climate impacts, including increased temperatures,
accelerated ice and snow melt, and disruptions in
precipitation patterns. BC is also a constituent of
fine particles (PM25) and is therefore associated with
an array of respiratory and cardiovascular health
impacts. This makes it ripe for emissions reduction
approaches that bring both climate and public health
benefits.
Like many air pollutants, BC's atmospheric fate
is affected by a number of complex physical and
chemical processes that may enhance or attenuate
BC's warming impacts. Some of these atmospheric
processes are not yet completely understood,
making it challenging to represent them accurately
in climate models and to project future impacts.
Furthermore, BC is always co-emitted with other
pollutants, many of which have offsetting climate
impacts. Thus, BC must be studied in the context of
the total emissions mixture coming from particular
sources. In its 2007 Fourth Assessment Report, the
Intergovernmental Panel on Climate Change (IPCC)
noted that the climate effects of particles remained
"the dominant uncertainty" in estimating climate
impacts (IPCC, 2007). Since that time, additional
research has helped to reduce this uncertainty,
through inventory improvements, advances in
measurement technologies and methods, and
increasing sophistication in the representation
of particle atmospheric chemistry in climate
models. Thus, though important uncertainties
remain, substantial progress has been made in
understanding the role of BC and other particles
in climate processes. Recent work has clarified BC's
climate effects and the emissions control approaches
necessary to mitigate these impacts.
To further efforts to understand the role of BC in
climate change, on October 29, 2009, the United
States Congress established requirements for the
U.S. Environmental Protection Agency (EPA) to
conduct a BC study as part of /-/./?. 2996: Department
of the Interior, Environment, and Related Agencies
Appropriations Act, 2010. Specifically, the legislation
stated that:
"Not later than 18 months after the date of
enactment of this Act, the Administrator, in
consultation with other Federal agencies, shall
carry out and submit to Congress the results of a
study on domestic and international black carbon
emissions that shall include
• an inventory of the major sources of black
carbon,
• an assessment of the impacts of black carbon on
global and regional climate,
• an assessment of potential metrics and
approaches for quantifying the climatic effects
of black carbon emissions (including its radiative
forcing and warming effects) and comparing
those effects to the effects of carbon dioxide and
other greenhouse gases,
• an identification of the most cost-effective
approaches to reduce black carbon emissions,
and
• an analysis of the climatic effects and other
environmental and public health benefits of
those approaches."
To fulfill this charge, EPA has conducted an intensive
effort to compile, assess, and summarize available
scientific information on the current and future
impacts of BC, and to evaluate the effectiveness
of available BC mitigation approaches and
technologies for protecting climate, public health,
and the environment. The results are presented
in this Report to Congress on Black Carbon. This
report has been peer reviewed by a special panel of
experts appointed under the Council on Clean Air
Compliance Analysis, a Federal Advisory Committee
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Chapter 7
chartered under the Federal Advisory Committee
Act (FACA), 5 U.S.C., App 2. The Black Carbon
Review Panel concluded that "the draft report
is comprehensive and well-written; summarizes
relevant scientific literature; and successfully
convey[s] a wealth of complex information" (Pope,
2011). This final report reflects numerous additions
and improvements suggested by the peer review
panel. In addition, the final report reflects input
from other federal agencies, many of which have a
great deal of expertise in this area and a range of
programs relevant to BC. As requested by Congress,
EPA has consulted with these other federal agencies
while developing this report, and has incorporated
material supplied by the agencies into key sections
of the Report, including chapters on emissions
inventories, health and climate science, and
sector-specific mitigation options.
1.1 Key Questions Addressed in this
Report
In evaluating the climate impacts and mitigation
opportunities for BC, it is essential to recognize
from the outset that BC presents a different kind
of climate challenge than CO2 and other long-
lived GHGs. BC's short atmospheric lifetime (days
to weeks) and heterogeneous distribution around
the globe result in regionally concentrated climate
impacts. Thus, the location of emissions releases is
a critical determinant of BC's impacts, which is not
the case for long-lived and more homogeneously
distributed GHGs like CO2. The composition of the
total emissions mixture is also key: since many co-
emitted pollutants such as sulfur dioxide, oxides of
nitrogen, and most organic carbon particles tend to
produce a cooling influence on climate, the amount
of BC relative to these other constituents being
emitted from a source is important. Furthermore,
BC is linked to a whole variety of effects beyond
warming. These include the darkening of ice and
snow, which reduces reflectivity and accelerates
melting; changes in the formation and composition
of clouds, which affect precipitation; and impacts on
human health.
These key characteristics of BC give rise to some
important questions addressed in this Report,
including:
1. What is BC, and how does it lead to climate
warming?
2. What is the net effect of atmospheric BC on
global and regional temperature change in terms
of both magnitude and time scale?
3. What is known about the magnitude of BC's
effect on snow and ice, and its impacts on
precipitation?
4. What is known about BC's contribution to PM25-
related human health impacts and other, non-
climate environmental impacts?
5. What kind of real-world BC data exists from
monitoring networks and other observational
research?
6. How large are U.S. and international emissions
of BC currently, which sectors are the main
contributors, and how are emissions projected to
change in the future?
7. What is the potential value of BC reductions
as a component of a broader climate change
mitigation program, taking into account both
co-pollutant emissions reductions and the public
health co-benefits?
8. What specific considerations will determine
preferred mitigation strategies in different
national and regional contexts?
9. What technologies and approaches are available
to address emissions from key sectors, and at
what cost?
lO.Which mitigation options represent potential
top-tier opportunities for key world regions,
including the United States?
In answering these questions, this Report focuses
on synthesizing available scientific information
about BC from peer-reviewed studies and other
technical assessments, describing current and
future emissions estimates, and summarizing
information on available mitigation technologies
and approaches, including their costs and relative
effectiveness. Given the number of recent studies
and the limited time available to complete this
Report, EPA did not seek to undertake extensive
new analysis (such as climate modeling of specific
BC mitigation strategies), but instead relied on
information available in the literature. The report
focuses on BC, where the bulk of scientific research
is available, but acknowledges the potentially
important role played by other light-absorbing
particles which are still subject to great uncertainty.
This Report also describes specific research and
technical information needed to provide a stronger
foundation for future decision-making regarding
appropriate and effective BC mitigation policies.
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Introduction
1.2 Other Recent Assessments of BC
Numerous international and intergovernmental
bodies, including the United Nations Environment
Programme (UNEP) and the World Meteorological
Organization (WMO), the Convention on Long
Range Transboundary Air Pollution (CLRTAP), and
the Arctic Council, have identified BC as a potentially
important piece of the climate puzzle. Each of these
bodies has recently prepared an assessment of BC
that included consideration of the impacts of BC on
climate, the potential benefits to climate of reducing
BC emissions, and/or the mitigation opportunities
that appear most promising. These assessments
have identified a number of additional actions—from
improvements in inventories to evaluation of specific
mitigation opportunities—that could be taken
to help gather further information about BC and
address emissions from key sectors.
In the Integrated Assessment of Black Carbon and
Tropospheric Ozone (UNEP and WMO, 2011a), UNEP/
WMO conclude that BC mitigation may offer near-
term climate benefits. This study was designed to
assess the role of BC and ozone in climate and air
quality, and to recommend mitigation measures
that could be expected to provide benefits in both
the climate and air quality arenas. Out of roughly
two thousand potential mitigation measures,
the UNEP/WMO analysis has identified a small
subset of measures as providing the largest
mitigation potential. The Assessment finds that full
implementation of the targeted measures (which
included methane reductions for ozone mitigation,
as well as BC reductions) could greatly reduce global
mean warming rates over the next few decades.
Specifically, the analysis suggests that warming
anticipated to occur during the 2030s based on
emissions projections could be reduced by half
through application of these BC and methane
measures. In contrast, even a fairly aggressive
strategy to reduce CO2 would do little to mitigate
warming over the next 20-30 years. The UNEP/
WMO Assessment concludes that while CO2 measures
clearly are the key to mitigating long-term climate
change out to 2100, BC and methane measures could
reduce warming and slow the rate of change in the
next two decades. The Assessment also recognizes
the substantial benefits to air quality, human health,
and world food supplies that would result from
reductions in BC and tropospheric ozone.
The CLRTAP Ad-hoc Expert Group on Black Carbon
and the Arctic Council Task Force on Short-Lived
Climate Forcers focused mainly on identifying
high-priority mitigation options and the need
for supporting information, such as national BC
emissions inventories. These groups did not conduct
independent scientific assessments; rather, after a
review of existing scientific literature, they concluded
that current evidence suggests that BC plays an
important role in near-term climate change. The
CLRTAP Ad-hoc Expert Group was co-chaired by
the United States and Norway. In its final report
presented to the Convention's Executive Body in
December 2010, the Expert Group highlighted key
findings, including:
• There is general scientific consensus that
mitigation of BC will lead to positive regional
impacts by reducing BC deposition in areas with
snow and ice.
• There is virtual certainty that reducing primary
PM will benefit public health.
• The Arctic, as well as alpine regions, may benefit
more than other regions from reducing emissions
of BC.
• Climate processes unique to the Arctic have
significant effects that extend globally, so action
must be taken in the very near term to reduce the
rate of warming.
• Impacts on the Arctic and alpine areas will vary
by country, but all countries will benefit from
local emissions reductions of BC and other
co-emitted pollutants.
The Expert Group concluded that because of the
public health benefits of reducing BC, as well as
the location of the countries across the Convention
regions in relation to the Arctic, the Executive Body
should consider taking additional measures to
reduce BC. The report included information about
key sectors and emphasized the need to develop
emissions inventories, ambient monitoring and
source measurements in an effort to improve the
understanding of adverse effects, efficacy of control
measures and the costs and benefits of abatement.
Based in part on the findings of the Expert Group,
the CLRTAP Executive Body decided to include
consideration of BC as a component of PM in
their ongoing process of revising the Gothenburg
Protocol.1 This decision marks the first time an
international agreement has attempted to address
the issue of short lived climate forcers in the context
of air pollution policy. Revisions to the Gothenburg
Protocol are expected to be completed in 2012.
The Arctic Council Task Force on Short-Lived Climate
Forcers was formed following the issuance of the
1 The Protocol to Abate Acidification, Eutrophication and Ground-level
Ozone was adopted in Gothenburg, Sweden in 1999. See http://www.
uniece,org/env/irtop/mu111 hi„ htuni.
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Chapter 7
Troms0 Declaration at the Arctic Council Ministerial
in April 2009.2 This declaration formally noted the
role that short-lived forcers may play in Arctic
climate change, and recognized that reductions of
emissions of these compounds and their precursors
have the potential to slow the rate of Arctic snow,
sea ice, and sheet ice melting in the near term. The
Task Force, which is being co-chaired by the United
States and Norway, is charged with identifying
existing and new measures to reduce emissions of
short-lived climate forcers (BC, ozone and methane)
and recommending further immediate actions that
can be taken. The Task Force focused its initial efforts
on BC, and presented a menu of mitigation options
to Arctic Nations at the Arctic Council Ministerial in
May of 2011 (Arctic Council, 2011). The key findings
of this Task Force include:
• Addressing short-lived climate forcers such as
black carbon, methane and ozone offers unique
opportunities to slow Arctic warming in the near
term.
• Black carbon emitted both within and outside of
the Arctic region contributes to Arctic warming.
Per unit of emissions, sources within Arctic
Council nations generally have a greater impact.
• Controls on black carbon sources that reduce
human exposure to particulate pollution improve
health, and in that regard many measures can be
considered no regrets.
• To maximize climate benefits, particulate matter
control programs should aim to achieve maximum
black carbon reductions.
In making these recommendations, the Task Force
recognized that not all measures to control ambient
PM necessarily reduce BC, and encouraged countries
to consider BC-focused strategies as a complement
to existing PM control programs in order to ensure
climate as well as health and environmental benefits.
Based on its findings, the Task Force recommended
that Arctic Council nations take action to reduce BC,
and laid out a menu of specific mitigation options
in key sectors such as land-based transportation,
stationary diesel engines, residential wood
combustion, agricultural and forest burning, and
shipping. The Task Force has also encouraged Arctic
2 The Arctic Council comprises the eight member states with land above
the Arctic Circle (Canada, Denmark including Greenland and the Faroe
Islands, Finland, Iceland, Norway, Russian Federation, Sweden, and
the U.S.), six permanent participants representing indigenous peoples
resident in those member states, andanumberof observers. The Council
does not have legally-binding authority over its members, but rather
promotes cooperation, coordination, and interaction regarding common
Arctic issues.
nations to develop and share domestic inventories
of BC emissions, which can be used to further
define—and refine—global inventories.
The Task Force is collaborating with the Arctic
Monitoring and Assessment Programme (AMAP)
working group, which established an Expert Group
on Short-Lived Climate Forcers to prepare a detailed
technical report, The Impact of Black Carbon on
Arctic Climate. This report (Quinn et al., 2011)
contains climate modeling results highlighting the
significance of BC emissions sources from different
sectors and different regions for the Arctic climate.
The AMAP report indicates that emissions both
from within the Arctic and from the rest of the world
affect Arctic warming, with sources near to or within
the Arctic having particularly significant impacts per
unit of emissions.
These concurrent international assessments
strongly suggest that reducing BC emissions
will slow the rate of warming and provide other
near-term benefits to climate, as well as protect
public health. The analyses conducted in support
of these assessments provide useful information to
clarify BC's role in climate change, the impact of key
emissions source categories, and the applicability
of different mitigation options. This Report to
Congress on Black Carbon builds upon these efforts,
summarizing and incorporating their key findings
as appropriate. Since all of the efforts mentioned
above have been conducted by international
bodies with a focus outside the United States,
readers are encouraged to read their final reports
and recommendations as additional sources of
information.
1.3 Organization of this Report
This Report is organized into twelve chapters and
seven technical appendices. Each of the chapters
that follow this Introduction, and the appendices, are
described briefly below:
Chapter 2 describes how particles, including BC,
absorb and scatter light, and identifies the factors
that influence the direction and magnitude of
their effect on the Earth's climate. The chapter
defines "black carbon," describes how BC relates
to other types of particles, and discusses how
these substances affect climate. Next, the chapter
provides detailed information on the range of
direct and indirect impacts of BC on global and
regional climate. It summarizes available estimates
of BC's global and regional radiative forcing and
related temperature effects, snow and ice albedo
effects, cloud effects, and precipitation effects.
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Introduction
The chapter also considers how best to compare
the effects of BC to the impacts of other climate
forcers, particularly CO2. It evaluates the applicability
of traditional metrics developed for CO2 to BC, and
presents alternative metrics designed specifically for
evaluating the climate impacts of short-lived climate
forcers like BC.
Chapter 3 outlines EPA's current scientific
understanding of the health and non-climate
environmental effects of BC. This chapter discusses
the large body of scientific evidence regarding the
adverse human health impacts of PM25 in general,
and provides a summary of health research related
to BC as a component of the overall PM2 5 mix. It
also describes BC's role in visibility impairment and
ecological effects.
Chapter 4 provides a detailed look at BC emissions
inventories. The chapter characterizes current
(2005) U.S. emissions of BC by source category,
and provides detailed information regarding
emissions from sectors that are the most significant
contributors to U.S. emissions, such as mobile
sources, open biomass burning, and stationary
fossil fuel combustion. The chapter also provides
an overview of global and regional emissions
inventories for BC, and contrasts these global
inventories with more refined regional inventories
available for some areas, such as the United
States, China and India. Special attention is paid to
emissions near the Arctic. The chapter discusses the
transport of emissions from particular sources and
regions, and describes historic emissions trends.
Chapter S summarizes key findings from
observational data on BC. This includes data from
ambient air quality monitors, ice/snow cores, and
remote sensing. The chapter describes the existing
BC monitoring networks, and summarizes available
data regarding ambient levels in urban and rural
areas, both domestically and globally. The chapter
also describes trends in ambient BC concentrations.
Chapter 6 considers the potential climate and
human health benefits of BC emissions reductions.
The chapter describes the findings of existing
studies on the global and regional benefits of BC
mitigation, including specific strategies aimed at
reducing emissions from key sectors. The chapter
acknowledges the large remaining uncertainties
with regard to evaluating the climate benefits of BC
mitigation in some sectors, but notes that controls
on BC and co-emitted pollutants are generally
associated with significant public health benefits,
through reductions in PM25 and its precursors. The
chapter also discusses approaches for valuing health
and climate impacts.
Chapter 7 provides a framework for evaluating
mitigation options. The chapter describes the
different considerations, including benefits,
costs, technologies, and other factors that
affect decisionmaking, and the impact of these
considerations on choices among BC mitigation
options. It provides some illustrative examples
of which mitigation options might be preferred
depending on the weight policymakers assign to
different factors in the mitigation framework.
Chapters 8-11 describe existing control programs
and technologies that have been demonstrated to
be effective in reducing BC emissions from source
categories of regional and/or global importance.
These include Mobile Sources (Chapter 8),
Stationary Sources (Chapter 9), Residential
Heating and Cooking (Chapter 10), and Open
Biomass Burning (Chapter 11). For each sector,
the chapter recaps current and projected emissions
estimates (accounting for control programs currently
in place but not yet implemented), describes key
control technologies and other mitigation strategies
that can help control BC emissions from specific
source types, and provides available information
regarding control costs. The chapters also discuss
how alternative strategies, such as changes in
land-use policy or energy systems, could impact
emissions from the sectors. Control options, costs,
and known or potential barriers to mitigation are
described separately for U.S. domestic emissions
and international emissions. In some cases, there are
considerable differences in mitigation approaches,
cost, and feasibility between the United States and
other countries. Also, there are gaps in available
information on these factors for many sectors.
The conclusion, Chapter 12, focuses on identifying
important BC mitigation opportunities for the
United States and other world regions. Drawing on
earlier chapters and the findings of other recent
assessments, this chapter clarifies some of the key
mitigation options that can clearly be expected to
provide near-term climate and health benefits. The
chapter acknowledges the diversity of approaches
for BC mitigation and the need to tailor mitigation
strategies to specific national and local contexts. The
chapter also identifies key gaps in current scientific
understanding, and provides a list of high-priority
research needs. Additional research in these areas
is essential for improving the current scientific
understanding of the impact of BC and other light-
absorbing particles on climate, and for estimating
the full impact of mitigation approaches in different
sectors and regions on both climate and public
health. These research needs may stimulate further
work on BC by EPA and other organizations.
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Chapter 7
Appendix 1 provides further details regarding
alternative definitions of BC and other light-
absorbing particles, and the techniques and
instruments used for ambient monitoring and
measurement of BC.
Appendix 2 provides a detailed explanation of the
methods that are used to compile U.S. emissions
inventories for BC. It also further explores the
variety of global and non-U.S. regional emissions
inventories available and some of the key
differences among those inventories.
Appendix 3 summarizes the results of available
studies which have estimated the public health
benefits that might accrue from alternative BC
mitigation strategies, at either the global or
regional level.
Appendix 4 describes world-wide efforts to
reduce the sulfur content of diesel fuels, which
is an important prerequisite to reducing BC
emissions from mobile sources.
Appendix 5 provides a full list of the emissions
standards for different categories of mobile
sources in the United States, and the emissions
limits set under those standards.
Appendix 6 describes existing emissions
standards for heavy-duty diesel vehicles
internationally, and the anticipated schedule
for emissions reductions resulting from these
standards.
Appendix 7 discusses a variety of research
needs related to BC. Though the highest priority
research needs are discussed in Chapter 12, this
appendix provides more detail regarding specific
gaps in the currently available information on
BC that are important both from a scientific
perspective and for informing BC mitigation
decisions.
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Chapter 2
Black Carbon and Its Effects on
Climate
2.1 Summary of Key Messages
• Black carbon (BC) is the most strongly light-
absorbing component of particulate matter (PM),
and is formed by the incomplete combustion of
fossil fuels, biofuels, and biomass.
- BC can be defined specifically as a solid
form of mostly pure carbon that absorbs
solar radiation (light) at all wavelengths. BC
is the most effective form of PM, by mass,
at absorbing solar energy. BC is a major
component of "soot", a complex light-
absorbing mixture that also contains organic
carbon (OC).
- Other carbon-based PM may also be light-
absorbing, particularly brown carbon (BrC),
which is a class of OC compounds that absorb
light within the visible and ultraviolet range
of solar radiation and that can exist within the
same particles as BC. The net contribution of
BrC to climate is presently unknown.
• BC is always emitted with other particles and
gases, such as sulfur dioxide (SO2), nitrogen
oxides (NOJ, and OC. Some of these co-emitted
pollutants exert a cooling effect on climate.
Therefore, estimates of the net effect of BC
emissions sources on climate should include the
offsetting effects of these co-emitted pollutants.
• Atmospheric processes that occur after BC is
emitted, such as mixing, aging, and coating, can
also affect the net influence of BC on climate.
• The short atmospheric lifetime of BC (days to
weeks) and the mechanisms by which it affects
climate distinguish it from long-lived greenhouse
gases (GHGs) like carbon dioxide (CO2).
- Targeted strategies to reduce BC emissions
can be expected to provide climate responses
within the next several decades. In contrast,
reductions in GHG emissions will take longer
to influence atmospheric concentrations
and will have less impact on climate on a
short timescale, but deep reductions in GHG
emissions are necessary for limiting climate
change over the long-term.
- Emissions sources and ambient
concentrations of BC vary geographically
and temporally, resulting in climate effects
that are more regionally and seasonally
dependent than the effects of long-lived,
well-mixed GHGs. Likewise, mitigation
actions for BC will produce different climate
results depending on the region, season, and
emissions category.
BC influences climate through multiple
mechanisms:
- Direct effect BC absorbs both incoming and
outgoing radiation of all wavelengths, which
contributes to warming of the atmosphere
and dimming at the surface. In contrast,
GHGs mainly trap outgoing infrared radiation
from the Earth's surface.
- Snow/ice albedo effect BC deposited on
snow and ice darkens the surface and
decreases reflectivity (albedo), thereby
increasing absorption and accelerating
melting. GHGs do not directly affect the
Earth's albedo.
- Other effects: BC also alters the properties
and distribution of clouds, affecting
cloud reflectivity and lifetime ("indirect
effects"), stability ("semi-direct effect"), and
precipitation. These impacts are associated
with all ambient particles, but not GHGs.
The direct and snow/ice albedo effects of BC are
widely understood to lead to climate warming.
Based on the studies surveyed for this report, the
direct and snow/ice albedo effects of BC together
likely contribute more to current warming than
any GHG other than CO2 and methane (CH4).
The climate effects of BC via interaction with
clouds are more uncertain, and their net climate
influence is not yet clear.
Report to Congress on Black Carbon
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Chapter 2
- All aerosols (including BC) affect climate
indirectly by changing the reflectivity and
lifetime of clouds. The net indirect effect of
all aerosols combined is very uncertain but
is thought to have a net cooling influence.
The contribution of BC to this cooling has not
been quantified.
- BC has additional effects on clouds—including
changes to cloud stability and enhanced
precipitation from colder clouds—that can
lead to either warming or cooling.
- The net climate influence of these cloud
interaction effects of BC is not yet clear.
There is inconsistency among reported
observational and modeling results, and many
studies do not provide quantitative estimates
of cloud impacts.
The sign and magnitude of the net climate forcing
from BC emissions are not fully known at present,
largely due to remaining uncertainties regarding
the effects of BC on clouds. Though most
estimates indicate that BC has a net warming
effect, a net cooling influence cannot be ruled
out. Further research and quantitative assessment
are needed to reduce remaining uncertainties.
Regional climate impacts of BC are highly variable,
and sensitive regions such as the Arctic and
the Himalayas are particularly vulnerable to the
warming and melting effects of BC. Estimates of
snow and ice albedo forcing in key regions also
exceed global averages.
BC also contributes to the formation of
Atmospheric Brown Clouds (ABCs) and resultant
changes in the pattern and intensity of
precipitation.
Due in large part to the difference in lifetime
between BC and CO2, the relative weight given to
BC as compared to CO2 (or other climate forcers)
in terms of its impact on climate is very sensitive
to the formulation of the metric used to make the
comparison.
There is currently no single metric that is widely
accepted by the science and research community
for this purpose.
There are several metrics that have been applied
to the well-mixed GHGs with respect to different
types of impacts, especially the global warming
potential (GWP) and global temperature potential
(GTP). These metrics can be applied to BC, but
with difficulty due to important differences
between BC and GHGs. Recently, new metrics
designed specifically for short-lived climate
forcers like BC have been developed, including
the specific forcing pulse (SFP) and the surface
temperature response per unit continuous
emission (STRE).
• There is significant controversy regarding the
use of metrics for direct comparisons between
the long-lived GHGs and the short-lived particles
for policy purposes; however, these comparisons
are less controversial when used for illustrative
purposes.
- There are a number of factors that should
be considered when deciding which metric
to use, or whether comparisons between BC
and CO2 are useful given a particular policy
question. These include: the time scale (e.g.,
20 years, 100 years, or more), the nature of
the impact (radiative forcing, temperature,
or more holistic damages), the inclusion of
different processes (indirect effects, snow
albedo changes, co-emissions), and whether
sources and impacts should be calculated
regionally or globally.
- If the primary goal is reducing long-term
change, then a metric like a 100-year GWP
or GTP would be more appropriate. If the
rate of near-term climate change and near-
term damages to sensitive regions like the
Arctic are also a consideration, there is
no single existing metric that adequately
weights impacts over both time periods,
and a multi-metric approach may be more
appropriate than developing a single metric
that attempts to serve all purposes.
2.2 Introduction
There is a general consensus within the scientific
community that BC is contributing to climate change
at both the global and regional levels. Like CO2, BC
is produced through the burning of carbon-based
fuels, including fossil fuels, biofuels and biomass.
BC is part of the mix of PM released during the
incomplete combustion of these fuels. BC influences
climate by absorbing sunlight when suspended in
the atmosphere or when deposited on the Earth's
surface. The energy absorbed by BC is then released
as heat and contributes to atmospheric warming
and the accelerated melting of ice and snow. In
addition, BC is capable of altering other atmospheric
processes, such as cloud formation and evaporation,
and precipitation patterns.
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Black Carbon and Its Effects on Climate
The strong absorption, short atmospheric lifetime,
and other characteristics of BC make its impacts
on climate different from those of long-lived GHGs
like CO2 (see Figure 2-1). Because BC is involved
in complex atmospheric physical and chemical
processes, it is difficult to disentangle all associated
impacts and to evaluate its net effect on climate. In
addition, the combustion processes that produce BC
also produce other pollutants, such as SO2, NOX, and
OC. Since many of these compounds have a cooling
effect, BC's impacts are mixed with—and sometimes
offset by—these co-emitted substances. This must
be considered when evaluating the net effect of
emissions sources.
This chapter focuses on how and to what extent
BC influences the Earth's climate. Specifically, this
chapter discusses approaches for defining BC
and other light-absorbing particles, highlights the
differences between BC and GHGs, and addresses
the role of co-emitted pollutants. Further, this
chapter summarizes recent scientific findings
regarding the processes by which BC affects climate
and the magnitude of BC's impacts on global and
Reflecting
Particles
,»•••»• •.
• -..
* I V
Black Carbon (BC)
©
Snow and/or Ice
Figure 2-1. Effects of BC on Climate, as Compared to GHGs. (Source: U.S. EPA)
1. Sunlight that penetrates to the Earth's surface reflects off bright surfaces, especially snow and ice.
2. Clean clouds and non-light-absorbing (transparent) particles scatter or reflect sunlight, reducing the amount of solar
energy that is absorbed by the surface.
3. BC suspended in the atmosphere absorbs some incoming solar radiation, heating the atmosphere.
4. Clouds containing BC inclusions in drops and BC interstitially between drops can absorb some incoming solar radiation,
reducing the quantity that is reflected. Clouds warmed by the absorbed energy have shorter atmospheric lifetimes and
may be less likely to precipitate compared to clean clouds.
5. BC deposited on snow and/or ice absorbs some of the sunlight that would ordinarily be reflected by clean snow/ice, and
increases the rate of melting.
6. Most solar radiation is absorbed by the Earth's surface and warms it. Part of the absorbed energy is converted into
infrared radiation that is emitted into the atmosphere and back into space.
7. Most of this infrared radiation passes through the atmosphere, but some is absorbed by GHG molecules like CO2,
methane, ozone and others. These gases re-emit the absorbed radiation, with half returning to the Earth's surface. This
GHG effect warms the Earth's surface and the lower atmosphere.
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Chapter 2
Figure 2-2. BC Images, (a) High resolution transmission electron microscopy (TEM) image of a BC spherule
(Posfai and Buseck, 2010). (b)TEM image of a representative soot particle. Freshly emitted soot particles are
aggregates of soot spherules (Alexander et al., 2008).
regional climate, highlighting the effect of BC in
sensitive regions such as the Arctic and other snow-
and ice-covered regions. The chapter discusses the
significant remaining uncertainties about BC's effects
on climate, and the need for further research in key
areas. The final section of this chapter introduces
several metrics that can be used to quantify the
climate impacts of BC and other pollutants (such as
CO2 and CH4) relative to a common baseline. The
section highlights the fact that there is no one "best"
metric for comparing BC to other pollutants and
that the utility of each metric depends on the policy
objective.
2.3 Defining Black Carbon and Other
Light-Absorbing PM
All PM in the atmosphere can affect the Earth's
climate by absorbing and scattering light. Sunlight
absorbed by PM increases the energy in the
Earth's climate system, leading to climate warming.
Conversely, light scattered by PM generally leads to
increased reflection of light back to space, leading
to climate cooling (Charlson, 1992; Moosmuller et
al., 2009; Seinfeld and Pandis, 2006; Forster et al.,
2007). Carbonaceous PM, a class of material found in
primary and secondary particles, has typically been
divided into two classes: BC and OC (see text box
on "Terminology"). Neither BC nor OC has a precise
chemical definition. The term BC generally includes
the solid forms of carbon emitted by incomplete
combustion while OC refers to the complex mixtures
of different carbon compounds found in both
primary and secondary carbonaceous particles.
Carbonaceous PM includes an array of organic
compounds that, along with BC, possess radiative
properties that fall along a continuum from light-
absorbing to light-scattering. Both BC and OC are
part of the broader category of suspended particles
and gases known as aerosols, all of which have light-
absorption and light-scattering properties.
In this report, BC is defined as the carbonaceous
component of PM that absorbs all wavelengths
of solar radiation.1 For this reason, among the
many possible forms of PM, BC absorbs the most
solar energy. Per unit of mass in the atmosphere,
BC can absorb a million times more energy than
CO2 (Bond and Sun, 2005), making it a significant
climate warming pollutant in regions affected by
combustion emissions.
BC forms during combustion, and is emitted when
there is insufficient oxygen and heat available for the
combustion process to burn the fuel completely (see
text box on "Products of Incomplete Combustion").
BC originates as tiny spherules, ranging in size from
0.001 to 0.005 micrometers (um), which aggregate
to form particles of larger sizes (0.1 to 1 u.m) (Figure
2-2). Particles in this range are similar in size to
the wavelengths emitted by the sun, making them
especially effective in scattering or absorbing these
wavelengths (Horvath, 1993). The characteristic
particle size range in which fresh BC is emitted also
makes it an important constituent of the ultrafine
(<100 nanometers (nm)) subclass of PM2.5.
'The spectrum of solar radiation striking Earth's atmosphere ranges
from high energy UV with wavelengths shorter than 280 nm down
to infrared radiation as long as 1000 nm. However, UV wavelengths
shorter than 280 nm are substantially absorbed by the stratosphere.
For the purposes of this discussion, the term "all wavelengths of
solar radiation" corresponds to the solar wavelengths present in the
troposphere (e.g., in the range 280 - 2500 nm).
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Black Carbon and Its Effects on Climate
Terminology
Black carbon (BC) is a solid form of mostly pure carbon
that absorbs solar radiation (light) at all wavelengths.
BC is the most effective form of PM, by mass, at
absorbing solar energy, and is produced by incomplete
combustion.
Organic carbon (OC) generally refers to the mix of
compounds containing carbon bound with other
elements like hydrogen or oxygen. OC may be a product
of incomplete combustion, or formed through the
oxidation of VOCs in the atmosphere.2 Both primary and
secondary OC possess radiative properties that fall along
a continuum from light-absorbing to light-scattering.
Brown carbon (BrC) refers to a class of OC compounds
that absorb ultraviolet (UV) and visible solar radiation.
Like BC, BrC is a product of incomplete combustion.3
Carbonaceous PM includes BC and OC. Primary
combustion particles are largely composed of these
materials.
Light absorbing carbon (LAC) consists of BC plus BrC.
Soot, a complex mixture of mostly BC and OC, is the
primary light-absorbing pollutant emitted by the
incomplete combustion of fossil fuels, biofuels, and
biomass.
BC is emitted directly from sources, making it a form
of primary PM. This distinguishes it from secondary
PM such as sulfates, nitrates and some forms of OC
that are formed in the atmosphere from gaseous
precursors like SO2, NOX and volatile organic
compounds (VOCs).
When BC is emitted directly from sources as a result
of the incomplete combustion of fossil fuels, biofuels
and biomass, it is part of a complex particle mixture
called soot which primarily consists of BC and OC.
This mixture is the light-absorbing component of
these air pollution emissions.
Soot mixtures can vary in composition, having
different ratios of OC to BC,2 and usually include
inorganic materials such as metals and sulfates. For
example, the average OCBC ratio among global
sources of diesel exhaust is approximately 1:1.
For biofuel burning, the ratio is approximately 4:1
and for biomass burning it is approximately 9:1
2When referring to emissions and measurements, OC denotes
the total carbon associated with the organic compounds, while
organic mass (OM) refers to the mass of the entire carbonaceous
material, including hydrogen and oxygen. Similarly, measurements
and emissions reported as elemental carbon (EC) denote the non-
organic, refractory portion of the total carbon and is an indicator for
BC. For more details, see Chapter 5 and Appendix 1.
(Lamarque et al., 2010). As expected, very dark soot
indicates the presence of low OCBC ratios. As the
OC fraction begins to dominate, the color of the
soot mixture shifts to brown and yellow. A brown
soot sample is dominated by a form of OC known,
as might be expected, as "brown carbon" (BrC).3 BrC,
another product of incomplete combustion, absorbs
portions of the visible spectrum, but is less effective
in capturing solar energy than BC (Alexander et al.,
2008; Novakov and Corrigan, 1995b). The mixture
shifts in color toward yellow when the emissions
source is no longer producing BC and BrC. Yellow
carbon, another form of OC, is also able to absorb
visible radiation, but to a lesser extent than
BrC (Bond, 2001; Gelencser, 2004; Andreae and
Gelencser, 2006). Figure 2-3 illustrates the variance
in soot composition resulting from different fuels
and stages of fuel combustion. The stages of fuel
combustion responsible for producing BC and the
various forms of OC observed in soot are described
in the text box on this page.
Figure 2-3. Representative Examples of Filter
Samples Collected from Different Sources,
including: (a) Smoldering Biomass, (b) Flaming
Biomass, and (c) Diesel Exhaust. (Photo courtesy of
Desert Research Institute)
In general, light absorption by carbonaceous PM can
be described as a continuum from light-absorbing
to light-scattering with BC at one end, most OC at
the other, and BrC occupying the partially absorbing
3 During solid fuel combustion, BrC forms during the preheating
(pyrolysis) phase, and during both flaming and smoldering
combustion. The light-colored smoke characteristic of the pyrolosis
and smoldering combustion phases is primarily OC, including both
BrC and other forms of OC, and does not include soot. Secondary
BrC can also form during reactions, similar to polymerization, that
take place in primary particles as emissions plumes age. BrC of this
type is known as "humic-like substances" (HULIS).
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Chapter 2
(a)
(b)
c
g
rt
T3
_rt
O
00
0)
Incoming Solar Radiation
1.6
1.2
0.8
0.4
500
^visible* ]•*•
1000
A, [nm]
^^—^^— infrared
1500
2000
Absorption by BC
Absorption by BrC
(light-absorbing OC)
QL
tu
E*.
CD
cr
10
q
•
o
•3
n>
500
1 000
1500
2000
, [nm]
(C)
O
V>
rt
T3
_rt
O
I/I
cu
Incoming Solar Radiation
Absorption by BC
Absorption by BrC
(light-absorbing OC)
73
500 1000 1500
A.[nm]
/4|* visible* | < infrared —
2000
Figure 2-4. Light Absorption by BC, BrC, and Ambient Mixtures. (Source: U.S. EPA)
a. The radiation wavelengths emitted by the Sun that reach the Earth begin around 280 nm (UV-B), peak in the mid-visible
range, and reach out past 2000 nm (Infrared). The shorter the wavelength, the higher its energy.
b. The extent to which BC and BrC absorb solar radiation depends upon the wavelength of incom ing light. This plot shows
idealized examples of those dependencies, assuming that both forms of carbon absorb to the same extent at 280 nm. BC
is more effective in absorbing solar energy across the entire solar spectrum than any form of BrC. BrC increasingly declines
in its capacity to absorbing longer wavelengths as the mixture moves from larger light-absorbing compounds to smaller
compounds, as indicated by the change in color from dark brown to yellow.
c. When the light absorption curves for BC and BrC are superimposed upon the solar spectrum, the significance of the
different absorption efficiencies between BC and BrC becomes evident. BC will, all else being equal, absorb more total
solar radiation than BrC. In practice, the mass ratios of BC and BrC, along with the specific composition of the BrC mixture,
determine the degree to which each form of carbon absorbs the solar energy penetrating the emissions plume.
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Black Carbon and Its Effects on Climate
Products of Incomplete Combustion
Most combustion occurring on Earth (both anthropogenic and natural) involves carbon-based fuels, including fossil fuels
(e.g., coal, oil, and natural gas), biomass (e.g., wood and crop residues), and biofuels (e.g., ethanol). Complete combustion
of a carbon-based fuel means all carbon has been converted to C02. Once ignited, the fuel must be well mixed with
oxygen at a sustained high temperature for this to occur. Incomplete combustion emits various materials in both gas and
particle form, depending on the combustion conditions (e.g., oxygen availability, flame temperatures, and fuel moisture)
and the type of fuel burned (e.g., gas, liquid, or solid). The PM emissions are generically known as soot.
BC is formed during the flaming phase of the combustion process. The quantity of BC emitted depends largely on
combustion conditions. If there is sufficient oxygen and high temperatures, the soot will be completely oxidized, and
BC emissions will be minimal.To increase fuel efficiency and reduce soot emissions, closed combustion systems (e.g.,
furnaces, combustors, reactors, boilers, and engines) are engineered to increase the mixing of air with the fuel and are
insulated to ensure temperatures remain high. Open and uncontrolled burning produces large quantities of BC because
oxygen availability and temperatures within the fire can vary widely.
The form of the fuel also influences the likelihood of complete combustion:
• Gas phase fuels (e.g., natural gas) can be readily mixed with oxygen, which reduces the emission of carbonaceous
particles.
• Liquid fuels (e.g., gasoline) generally must vaporize in order to fuel flaming combustion. If a liquid fuel contains
heavy oils, vaporization and thorough mixing with oxygen are difficult to achieve. The heavy black smoke emitted
by some marine vessels (which burn a sludge-like grade of oil known as "bunker fuel") is evidence of substantial BC
emissions.
• Solid fuels (e.g., wood) require preheating and then ignition before flaming combustion can occur. High fuel
moisture can suppress full flaming combustion, contributing to the formation of BrC particles as well as BC (Graber
and Rudich, 2006; Posfai et al., 2004; Alexander et al., 2008).
Thermal breakdown of high molecular weight fuels, known as pyrolysis, produces a wide array of BrC compounds.
When sustained, pyrolysis converts solid fuels such as coal and biomass into char, while releasing volatile gases that
can fuel flaming combustion.There is also a non-flaming process known as smoldering that is a slower, coolerform of
combustion which occurs as oxygen directly attacks the surface of heated solid fuel.The smoke that appears is light-
colored, consisting of a variety of organic particles composed of BrC. BC does not form under these conditions, since
temperatures are too low to sustain flaming combustion. During open or uncontrolled burning of solid fuels, all stages of
the burning process—pyrolysis, smoldering, and flaming combustion—occur simultaneously, in different parts of the fuel
pile, resulting in emissions of both BC and BrC.
middle ground.4 The characteristic light absorption
spectra vary significantly among individual BrC
compounds, but are almost entirely limited to the UV
to visible portion of the solar spectrum (Jacobson,
1999). Mixtures of these compounds range in color
from yellow to brown, roughly corresponding to
the average molecular weight of the light-absorbing
compounds present (see Figure 2-4). Emissions
dominated by smaller BrC compounds will appear
yellow, while plumes containing high concentrations
4Particles containing iron and other calcium, aluminum, and
potassium oxides also absorb light. Like BrC, metal oxides are
very effective, more so than BC, at absorbing light at shorter
wavelengths. Some metal oxides are derived from heavy fuel
sources such as residual fuel oil (Huffman et al., 2000). High
concentrations of such particles can result from windblown dust
and may be significant during dust and sand storms that occur in
Africa, China, and the Middle East. These fine particle constituents
can travel long distances and may contribute to a positive radiative
forcing to a limited degree (Prospero et al., 2010; Liu et al., 2008a).
of heavy polycyclic aromatic hydrocarbons (PAHs)
will appear darker brown.
BC and different mixtures of BrC show different
patterns of light absorption versus wavelength.
Light absorption by BC tends to decline more slowly
with increasing wavelength, while the falloff in
absorption by BrC is always faster than that of BC,
tending to vary depending on the composition of
the BrC mixture. Absorption by the aforementioned
yellow BrC mixtures, dominated by lower molecular
weight compounds, falls off very quickly. Dark
brown BrC mixtures containing large PAHs or HULIS
(see footnote 3) continue absorbing to a significant
degree across the solar spectrum.
BrC typically accompanies BC in soot particles.
However, independent BrC particles may form
in the uncontrolled burning conditions typical of
biomass burning, or during inefficient combustion
of biofuels. These independent particles, labeled "tar
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Chapter 2
Figure 2-5. TEM Image of a BrC Particle.
These particles are referred to as "tar
balls" or "carbon spheres" in the literature.
(Alexander et al., 2008)
Figure 2-6. Coarse Urban PM (Diameter
> 2.5 microns) with a Black Surface Coating,
shown on a Filter from a Harvard Impactor
Run Approximately 100 feet Above Street
Level in Boston. (Photo courtesy of NESCAUM)
balls" or "carbon spheres," range in diameter from
0.03 and 0.5 urn (Posfai et al., 2004; Chakrabarty et
al., 2010; Adachi and Buseck, 2008). Alexander et al.
(2008), in their study of East Asian pollution plumes,
found numerous examples of these particles (see
Figure 2-5). They noted that, on the basis of their
statistical analysis, these particles are very abundant.
Fuel type and burning conditions determine
the quantity of BrC produced from a particular
combustion source. Some sources, such as open
biomass burning, can produce substantially more
BrC than BC. On the basis of quantity alone, BrC may
lead to greater total solar energy absorbed than BC
for those sources, despite the fact that BrC absorbs
less energy than BC per unit mass.
Until recently, most measurements of light absorbing
carbonaceous PM focused on BC, classifying all other
carbon as OC. Growing awareness of the presence
of light-absorbing BrC in biomass emissions has
prompted the recent suggestion from the scientific
community that absorption by BrC should be
explicitly included in the evaluation of the role of LAC
in climate warming, by accounting for emissions of
BrC, along its known radiative properties, in climate
modeling studies. Unfortunately, those data do not
yet exist.
Combustion emissions, depending upon the purity
of the fuel and burning conditions, can also contain
a number of inorganic pollutants. For example,
combustion of high-sulfur coal is a well-known
source of SO2 and sulfuric acid emissions. Mineral
salts, such as potassium chloride, are emitted during
biomass burning due to the presence of potassium
in wood and other plant materials. Localized regions
of very high heat within a combustion mixture
produce NOX, which further reacts to form nitrates.
These co-pollutants in a combustion plume can
subsequently form light-scattering particles. The
light scattered by these particles may offset the
warming effect due to the light absorbed by BC and
BrC in the emissions plume. This effect is discussed
further later in this chapter in sections 2.5 and
2.6.1.5.
The focus of the discussion up to this point has been
on the light absorbing properties of carbonaceous
PM (i.e., of the classes of material that appear in
particle form in the atmosphere). The chemical
composition and physical structure of individual
particles are also factors determining the overall
radiative properties of an emissions plume. Size
and morphology influence the efficiency of light
absorption by a particle. Particle size and chemical
composition change as a fresh emissions plume
begins mixing with the ambient atmosphere. Light
absorption by BC can be enhanced by 30% to 100%
when chemical processing in the atmosphere creates
a transparent coating5 on the surface of the particle
(Fuller et al., 1999; Shiraiwa et al., 2009; Bond et al.,
2006a). These effects and the role they play in the
overall radiative effect of a combustion plume on
regional climate are discussed in section 2.5.
5Coatings effective in enhancing light absorption by BC include
sulfuric acid from the oxidation of SO2, water from cloud-processing
and secondary OC that does not absorb solar radiation.
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Black Carbon and Its Effects on Climate
There is also a significant BC component in (or rather
on) coarse particles (PM1025 and larger), especially in
urban areas where coarse-mode particles (such as
from re-entrained road dust) are often coated with
BC, as shown in Figure 2-6. This coarse urban PM is
black, not earth-colored, and likely results from a
BC surface coating of coarse mode particles, rather
than from a uniform BC composition. While coarse
particles will have a very limited effect on climate,
they represent a means of human exposure to BC.
Chapter 3 discusses the currently understood effects
of coarse particles and BC on human health.
BC has been studied as a component of soils and
sediments, in addition to its role in air pollution
and climate. It plays an important role in various
biological, geochemical processes, and has been
used as a marker for local vegetation fire histories
(Schmidt and Noack, 2000). BC has also been the
subject of intense study in the combustion science
and engineering fields (Frenklach, 2002). Given
the number of distinct scientific disciplines that
have studied BC, and the different chemical and
environmental contexts in which it appears, many
different measurement and estimation approaches
exist in the literature, each with their own operational
definitions of BC. In the atmospheric sciences alone,
the terms "graphitic carbon", "apparent elemental
carbon" (ECa), "equivalent black carbon" (BCe),
"light absorbing carbon", "carbon black", "soot" and
"black smoke" are used interchangeably with or as
surrogates for BC materials (Bond and Bergstrom,
2006; Andreae and Gelencser, 2006).
Although these commonly used terms are not
strictly equivalent, we believe that the validity of
our analyses and conclusions are not materially
compromised by our adopting the convention of
using surrogate measurements for BC and soot.
EPA has traditionally used surrogate or indicator
measurements for many pollutants, including
PM2.5 whose current regulatory characterization
by EPA is based on stable, historically available,
consistent and reproducible measurements (Watson
et al., 1995). For this report, we believe that BC
or elemental carbon (EC) measurements are the
best available indicators of BC and soot as these
particles are only directly emitted from incomplete
combustion, whereas OC can be derived from
several sources (e.g., pollens, spores, condensed
vapors, secondary aerosols). Nevertheless, the
most commonly used measurements may not fully
capture the light absorption by BrC, and thus current
emissions estimates (and observations on which
they are based) may underestimate the positive
radiative forcing associated with these particles.
The connections between highly correlated BC and
EC measurements and their physical properties are
discussed further in Chapter 5.
For purposes of regional air quality management
(e.g., human health studies related to air quality,
the evaluation of modeled estimates, and the
attribution of emissions to sources), BC is measured
as a constituent of ambient PM2.5, and is expressed
in units of mass. Moreover, BC emissions inventories,
as discussed in Chapter 4, are also generally
expressed in mass units (e.g., tons/year).
Thus, light-absorption measurements are often
converted into estimates of carbon mass. This
practice, however, may also contribute to some of
the uncertainty in reported ambient concentrations
and emissions estimates. Due to reliance on
these mass-based indicators, BC is frequently
labeled EC due to the long-standing use of carbon
measurement methods from which the air quality
and emissions estimates were derived. This issue is
also evident in health studies that sometimes make
a distinction between BC and the measurement on
which it is based.
These issues are discussed more fully in Chapter 5
and Appendix 1, which provide a brief description
of the various BC and LAC absorption- and
mass-related measurement approaches.
2.4 Key Attributes of BC and
Comparisons to GHGs
The net impact of BC on climate depends on a
number of other factors in addition to its powerful
light-absorption capacity. These include atmospheric
lifetime, the geographic location of emissions,
altitude, interactions with clouds, the presence of
co-emitted pollutants, and the influence of aging
and mixing processes in the atmosphere. In many
of these aspects, BC differs substantially from
long-lived GHGs, as summarized in Table 2-1. These
differences have implications for how BC influences
climate and the climate benefits of BC mitigation
as compared to CO2 mitigation. Each of these
dimensions is explored further below.
Particles in general have relatively short
atmospheric lifetimes in comparison to GHGs.
Particles of any type, including BC, are removed
from the atmosphere within days to weeks by
precipitation and/or dry deposition to surfaces.
This short atmospheric lifetime curtails their total
contribution to the Earth's energy balance, even for
those particles like BC that have strong absorptive
capacity. The efficiency with which particles are
removed is influenced by their size and chemical
Report to Congress on Black Carbon
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Chapter 2
Table 2-1. Comparison of BC to CO2 on the Basis of Key Properties that Influence the Climate. (Source: U.S. EPA)
Property BC CO2
Atmospheric lifetime
Distribution of atmospheric
concentrations
Direct radiative properties
Global mean radiative forcing
Cloud interactions
Surface albedo effects
Contribution to current global
warming
Dimming
Acidification/fertilization
Days to weeks
Highly variable both geographically and
temporally, correlating with emission sources
Absorbs all wavelengths of solar radiation
+0.34 to 1.0Wm2directforcinga
+0.05 W m 2 (snow/ice albedo forcing)15
± ? (cloud interactions)11
Net effect: uncertain, but likely warming
Multiple cloud interactions that can lead to
warming or cooling (typically cooling), as well as
effects on precipitation
Contributes to accelerated melting of snow/ice
and reduces reflectivity by darkening snow and
ice, enhancing climate warming
Likely third largest contributor (after C02 and
CH4), but large uncertainty11
Contributes to surface dimming
No ocean acidification/fertilization effects
Up to millennia6
Generally uniform across globe
Absorbs only thermal infrared radiation
+1.66 (±0.17) Wm2<"
Increases cloud droplet acidity
No direct surface albedo effects
Largest contributor
No direct effects on surface dimming
Main contributor to ocean acidification and
fertilization
a UNEP and WMO (2011a) estimate narrower central range of +0.3 to +0.6 W m"2.
b Some adjustment to this value may be appropriate to account for the greater warming efficacy of BC deposited on snow and ice. UNEP
and WMO (2011a) suggest a range of +0.05 to +0.25 W m 2 for BC snow/ice albedo forcing (adjusted for efficacy).
c Values are highly uncertain. The IPCC estimated that the cloud albedo effect of all aerosols combined was -0.7 W m 2, but did not include
other cloud effects and did not estimate the albedo effect of BC alone. UNEP and WMO (2011a) provided a central forcing estimate for
all of the cloud impacts of BC of-0.4 to +0.4 W m 2.
d Based on the IPCC forcing estimates (shown in Figure 2-10), the central estimate of the UNEP assessment (UNEP and WMO, 2011), and
this report's assessed range of BC forcing (from Figure 2-11), it is likely that the net BC forcing will be less than that of CH4 but it is
possible that BC might be the second largest contributor to warming, depending on uncertainties in the direct and indirect contributions
of BC to warming.
e The lifetime of CO2 is more complicated than for most other GHGs. The carbon in CO2 cycles between the atmosphere, oceans,
ecosystems, soil, and sediments. Carbon added to the carbon cycle is removed very slowly (over thousands of years) through processes
such as weathering and calcium carbonate formation (Archer et al., 2009). However, even if it is not removed from the carbon cycle,
carbon added to the atmosphere can also cycle to other media: approximately three quarters of the added carbon will, over a time scale
of decades or centuries, move out of the atmosphere into the ecosystem or oceans.
Forster etal. (2007).
composition. For example, atmospheric aging can
increase the size of a particle or alter its chemical
composition in a way that makes it an efficient
nucleus for cloud droplet formation, facilitating its
removal by precipitation.
By contrast, GHGs have longer atmospheric lifetimes.
This enables them to become well mixed in the
atmosphere and to continue to absorb energy over
many decades or centuries. Gases such as nitrous
oxide (N2O), CH4, or hydrofluorocarbons (MFCs)
have lifetimes that range from as short as a year
for some of the MFCs to as long as 50,000 years
for tetrafluoromethane (CF4), a perfluorocarbon
(Forster et al., 2007).6The carbon in CO2 cycles
between the atmosphere, oceans, ecosystems,
soil, and sediments; therefore, CO2 does not have
a single defined lifetime. Computer models have
indicated that about half of an emissions pulse of
6Several months are required in order for a gas to mix throughout a
hemisphere, and one to two years are required for a gas to become
well-mixed globally. Some gases, such as CH4, are well-mixed but
are included in the category of "short-lived climate forcer" because
a decadal lifetime has different implications for mitigation decisions
than the lifetimes of a century or more for many other GHGs. On
the other hand, ozone and water vapor have short lifetimes and,
like black carbon and other aerosols, are not well-mixed in the
atmosphere.
26
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Black Carbon and Its Effects on Climate
CO2 will disappear within 30 years, 30% within a
few centuries, and the last 20% may remain in the
atmosphere for thousands of years (Denman et al.,
2007).
BC's short atmospheric lifetime means that
atmospheric concentrations are highest near
significant emissions sources and during time
periods and seasons of emissions releases. This
high spatial and temporal variability affects
BC's impacts on climate. BC is a regional pollutant.
CO2 and other GHGs with lifetimes longer than a
year are global pollutants with relatively uniform
concentrations around the globe. It is generally
assumed that CO2 and other well-mixed GHGs have
essentially the same effect on climate regardless of
the location or season of emissions. The same is not
true for BC.
Geographic location and altitude are important
determinants of the impact of BC on climate.
Fine combustion particles including BC can be
transported up to thousands of miles from sources.
Particles have a greater effect on the net absorption
of solar radiation by the atmosphere when they are
emitted or transported over light-colored, reflective
(i.e., high "albedo") surfaces such as ice, snow, and
deserts. In the absence of PM, a high percentage
of sunlight would reflect off these surfaces and
return to space. Therefore, any absorption of either
incoming or reflected light by PM above these
surfaces is more likely to lead to warming than
absorption of light by PM above darker surfaces.
Even PM that is typically classified as reflecting
can darken these bright surfaces and contribute
to warming (Quinn et al., 2011). This mechanism
explains why studies have found the effects of
BC to be magnified in the Arctic and other alpine
regions, as discussed in sections 2.6.4 and 2.6.5.
In addition, the net radiative effects of BC can be
sensitive to altitude. A modeling study by Ban-
Weiss et al. (2011) suggests that while BC at low
altitudes (where most BC is indeed located) warms
the surface considerably, BC at stratospheric or
upper-tropospheric altitudes may decrease surface
temperature. In addition, as with particles suspended
above a bright desert or glacier, particles suspended
above bright cumulus clouds can absorb both
incoming and outgoing solar radiation, increasing
the net radiative effect of the light absorbing
particle. When suspended between cloud layers or
beneath a cloud, the particle may be shielded from
incoming light, therefore lessening its potential
radiative impact (Schulz et al., 2006).
Other key distinguishing features of BC include
the wide range of mechanisms through which it
influences climate and its association with other
adverse, non-climate related public health and
welfare effects. In addition to the direct radiative
forcing characteristic of both BC and GHGs, BC has
significant interactions with clouds that can result
in both warming and cooling effects. It can also
cause melting and warming via deposition to snow
and ice. BC and other particles are also directly
associated with a host of other environmental
effects, such as changes in precipitation patterns
and surface dimming. All of these effects are
discussed in greater detail later in this chapter.
GHGs, on the other hand, influence climate mainly
through direct radiative forcing effects. GHGs do not
directly interact with clouds, snow and ice, though
the warming of the atmosphere due to GHGs does
influence cloud formation, snow melt, and many
other climate properties. In addition, CO2 has a
fertilization effect on plants and an acidification
effect in the ocean, and CH4 emissions lead to
increased ozone concentrations and changes in the
lifetime of other atmospheric pollutants. Finally,
as a constituent of PM2.5, BC is directly linked to a
range of public health impacts (see Chapter 3). This,
too, distinguishes it from long-lived GHGs, which
affect public health and welfare primarily via climate
change effects.
An important implication of BC's strong absorptive
capacity, coupled with its short atmospheric
lifetime, is that when emissions of BC are reduced,
atmospheric concentrations of BC will decrease
immediately and the climate, in turn, will respond
relatively quickly. The potential for near-term
climate responses (within a decade) is one of the
strongest drivers of the current scientific interest in
BC. Mitigation efforts that reduce BC emissions can
halt the effects of BC on temperature, snow and ice,
and precipitation almost immediately. This means
that reductions of BC may have an immediate and
important benefit in slowing the near-term rate of
climate change, especially for vulnerable regions
such as the Arctic and the HKHT region. In contrast,
when long-lived GHG emissions are reduced,
the climate takes longer to respond because
atmospheric GHG concentrations—the result of
cumulative historic and present-day emissions—
remain relatively constant for longer periods (see,
for example, Figure 2-7). It is important to recognize,
however, that the short atmospheric lifetime of BC
also means that reductions in current BC emissions
will have much less impact on temperature many
Report to Congress on Black Carbon
27
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Chapter 2
E
HI |g
TO
O)
c
'o
0.5
-§.
_0
O
iO.06
CD
§0.04
I 0.02
Q.
,
-------
Black Carbon and Its Effects on Climate
3.0
2.5
6 2.0
Oi
oo
QJ
g1 1.5
fD
1.0
oi
Q-
Reference
CH, measures
CH4+BC Group 1^
measures
CH4 + allBC
measures
all BC measures
0.5
2010 2020 2030 2040 2050 2060 2070
3.0
2.5
2.0
1.5
1.0
0.5
2010 2020 2030 2040 2050 2060 2070
all BC measures
Figure 2-8. Projected Global Mean Temperatures under Various Scenarios Relative to the 1890-1910
Average. This figure is adapted from Figure 6.2 from the UNEP/WMO Assessment. The left panel shows a
reference scenario, and then a sequence of emissions reduction scenarios in which first a set of methane
reductions are implemented, then one set of BC measures, then a second set of BC measures, and then a
set of CO2 measures. In the right panel, the same reference scenario is shown, followed by a sequence of
emissions reduction scenarios in which first the same set of CO2 measures is applied, followed by some
additional GHG reductions (mainly methane), followed by the full set of reductions from the left panel.
More details about these measures can be found in the UNEP/WMO report. (Note that modeled effects of
all measures include reductions of co-emissions, including reductions in cooling aerosols, which in the case
of CO2 measures resulted in a slight warming effect in the early years.) (UNEPand WMO, 201 la)
accompanied by OC, including BrC and other
carbonaceous materials. In addition, an emissions
plume may contain water, inorganic potassium and
sodium salts, ammonium nitrate and sulfate, gaseous
constituents (e.g., SO2, NOX and VOCs), various
hazardous air pollutants (e.g., metals), and even soil
particles.
The absorptive properties of an emissions plume
from a specific source will depend on all of
the co-emitted pollutants, and on how these
constituents interact with one another and other
atmospheric constituents in the atmosphere. As
described above, BC is co-emitted with OC and/or
sulfate, nitrate and gaseous constituents (SO2, NOX,
and VOCs). Since OC and sulfate and nitrate particles
generally exert a net cooling influence, these
pollutants play an important role in determining
the net absorptive capacity of the emissions
plume. These other constituents, however, may be
emitted in greater volume than BC, counteracting
the warming influence of BC. Thus, estimating
the climate impact of BC quantitatively requires
accounting for the impact of these co-emitted
pollutants. Emissions from a single source can also
vary over time. For example, the flaming phase of a
wildfire produces much more BC than its smoldering
phase. Also, when diesel trucks are under load, they
produce more BC than during other parts of their
driving cycle. Total particle number also impacts
scattering and absorption: the more particles
present in a portion of the atmosphere, the greater
the probability that light rays will be scattered or
absorbed by some of these particles.
Emissions from particular sources are often
characterized in terms of their OC to BC ratio.
Sources whose emissions mixtures are richer in
BC relative to the amount of OC emitted (i.e., with
lower OCBC ratios) are more likely to contribute to
climate warming; therefore, mitigation measures
focusing on these sources are more likely to
produce climate benefits. These ratios are useful in
that they take the emissions mixture into account;
Report to Congress on Black Carbon
29
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Chapter 2
however, they rely on crude accounting methods
and cannot provide precise measures of a particular
source's climate impacts. A particular concern is
the common presumption that all OC is cooling,
when in fact some components (especially BrC) are
light-absorbing and may contribute to the warming
associated with an emissions mixture. (The use of
OCBC ratios is discussed further in Chapter 7.)
A fresh emissions plume contains particles of many
different chemical compositions. Atmospheric
scientists refer to an emissions plume with this
kind of high inter-particle chemical variability as
"externally mixed." The externally mixed plume,
however, undergoes rapid chemical and physical
transformations. The coagulation of particles,
assuming no secondary particles form through
atmospheric chemical reactions, reduces the
overall particle number. This process, combined
with thermodynamically-driven mixing processes
such as water condensation and the redistribution
of semi-volatile PM compounds, reduces the
differences in chemical composition among the
individual particles in the plume. Over time, the
chemical composition of the particles within a given
plume approaches uniformity. A plume is said to be
"internally mixed" when it is near this theoretical
end-point. In situ measurements indicate that
emissions become internally mixed within a few
hours (Moffet and Prather, 2009b).
BC is an insoluble material, thus a well-mixed or
"aged" emissions plume will contain particles where
BC appears as an inclusion in an otherwise liquid
particle. These particles are often described as
"coated" BC particles. As described in section 2.3,
common coatings include sulfuric acid, water, and
transparent OC. The degree of mixing, or more
specifically, the fraction of BC particles that are
coated, influences the absorptive properties of the
particle. Internal mixtures of particles that include BC
have been observed to absorb light more strongly
than pure BC alone, by 30 to 100%. Whether BC is
modeled as an externally or internally mixed particle
can have a large effect on resulting estimates of
radiative forcing (see section 2.6.1.2).
Emissions plumes from different sources interact
with each other as well as with the surrounding
atmosphere. As a combustion emissions plume
rises into the atmosphere, it is diluted by ambient
air (see Figure 2-9). The open atmosphere contains
9
• o
o
• W e° ® (I1
ra ^•o0^^0
?o* ® ® 1
o _* • •
o •
Atmospheric Transport,
Photo- and Thermochemical
Transformation
Particle Dynamics
Fresh
Combustion
Emissions
Human Exposure
& Surface Deposition O
Figure 2-9. Particle Transformation in the Atmosphere, from Point of Emission to Deposition. A variety of physical and
chemical processes contribute to changing the light-absorption capacity of a fresh plume. (Source: U.S. EPA)
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Black Carbon and Its Effects on Climate
a number of reactive gases and particles originating
from a wide variety of anthropogenic and natural
sources. Physical and chemical changes resulting
from coagulation, condensation, and other
photochemical and atmospheric processes can alter
the climate-forcing impact of a given emissions
plume (Lauer and Hendricks, 2006) (see Table 2-2).
These mixing and aging processes are complex.
However, excluding them from models of the climate
impacts of BC may yield incomplete or erroneous
estimates. For example, coating of a BC particle by
a clear (light-scattering) shell has been shown to
enhance light absorption because the shell acts
as a lens that directs more light toward the core
(Ackerman and Toon, 1981; Jacobson, 2000; Lack
and Cappa, 2010; Jacobson, 2001). Other authors
have also found that light-absorption by BC is
enhanced when BC particles are coated by sulfate
or other light scattering materials (Shiraiwa et al.,
2009; Sato et a I., 2003; Moffet and Prather, 2009a;
Bond et al., 2006a). Other atmospheric processes,
however, such as further chemical processing or
Table 2-2. Examples of Particle Types and Mixtures Present in Combustion Plumes. The size, shape, and
chemical composition of a particle or particle mixture determine its radiative properties.
• o
o
o
n
®
3®
O
%
®
Type Radiative Properties
Black carbona
Brown (or yellow) carbon'3
Non-absorbing carbonb
Nitrate"
Sulfatec
Black carbon coated with brown
or non-absorbing carbond
Black carbon associated with
sulfate or nitrate6
Cloud and fog dropletsf
Complex of several particles6
Mixed particle (cloud
processed)9
Absorbing (all solarwavelengths)
Absorbing (UV and some visible)
Scattering
Scattering
Scattering
Absorbing (enhanced by partial internal
reflection of solar radiation); fractionally
scattering
Absorbing plus some scattering
Scattering
Absorbing and scattering
Absorbing (enhanced by partial internal
reflection of solar radiation);fractionally
scattering
Fresh BC is produced primarily during flaming combustion.
Particles condense within a fresh combustion plume from pyrolytic BrC and yellow OC. Oxidation of anthropogenic and
biogenic VOCs produces non-light absorbing carbon particles, and may also produce BrC and yellow carbon PM.
Emitted directly as a byproduct of combustion, or formed through the oxidation of SOx or NOX.
In the exhaust gases of solid fuel fires, low volatility BrC and other organic com pounds can condense on BC particles. In
the ambient atmosphere, low volatility organics produced by oxidation of VOCs can also condense on BC.
Forms when high particle concentrations lead to the coagulation of multiple particles.
Forms by condensation of water vapor onto acidic organic (carbon-based) and inorganic particles.
Forms when complex particles undergo the humidification and drying cycles characteristic of cloud formation and
evaporation.
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Chapter 2
particle growth through coagulation, may off-set
this enhancement. Observations by Chan et al. (2011)
at a rural site in Ontario, where BC particles were
assumed to be coated, did not show enhanced light
absorption. Furthermore, coated particles are more
easily removed by cloud droplets and precipitation,
decreasing their atmospheric lifetime (Stier et al.,
2006).
2.6 Global and Regional Climate
Effects of Black Carbon
BC affects climate through both direct and indirect
mechanisms. The most extensively studied of these
mechanisms is radiative forcing, which is directly
linked to temperature change. Radiative forcing is
a measure of how a pollutant affects the balance
between incoming solar radiation and exiting
infrared radiation, generally calculated as a change
relative to preindustrial conditions defined at 1750.
A pollutant that increases the amount of energy in
the Earth's climate system is said to exert "positive
radiative forcing," which leads to warming.8 In
contrast, a pollutant that exerts negative radiative
forcing reduces the amount of energy in the Earth's
system and leads to cooling. The net radiative
impact of a pollutant since preindustrial times
can be averaged over the Earth's surface and is
expressed in Watts per square meter (W rrr2). Global
average radiative forcing is a useful index because
in general it is related linearly to the global mean
temperature at the surface (Forster et al., 2007) and
is approximately additive across pollutants. Radiative
forcing also provides a consistent measure for
comparing the effects of past and projected future
emissions. As a result, it has become a standard
measure for organizations like the IPCC and the U.S.
Global Change Research Program (National Research
Council of the National Academies, 2005).
In addition to radiative forcing, BC is associated with
other effects including surface dimming and changes
in precipitation patterns. While not directly linked
to net global temperature change, these effects
also have important global and regional climate
implications. Each of these effects is discussed in
greater detail later in this section.
This section mainly addresses the climate impacts
of BC, GHGs, and other substances based on
the radiative forcing resulting from the change
in concentrations of these substances since
8 In general, radiative forcing in this document refers to "top-
of-the-atmosphere" radiative forcing unless otherwise specified.
Measuring at the top of the atmosphere (TOA) (in this context,
between the troposphere and the stratosphere) is the best location
for determining net energy balance.
Radiative Forcing
Radiative forcing: The change in the energy balance
between incoming solar radiation and exiting infrared
radiation, typically measured in watts per square
meter (W rrr2), due to a change in concentration
(generally the change since preindustrial conditions
in 1750). Positive radiative forcing tends to warm the
surface of the Earth, while negative forcing generally
leads to cooling.
preindustrial times. There are, however, other ways
to address climate impacts, and the decision of
which approach to use depends on at least two key
issues.
The first issue is that an analysis based on changes
in concentrations since preindustrial times is in
some ways a historical measure. An alternative
approach would be to analyze the effect of current
day emissions on radiative forcing into the future
[e.g., Figure 2-19 in this report, Figure 2.22 in
Forster et al. (2007), or the analysis by Shindell et al.
(2009)]. The benefit of such an approach is that an
emissions-based analysis is more policy-relevant,
because policies directly control emissions and not
concentrations. There are several disadvantages to
using this approach. First of all, fewer studies have
used this approach. Second, the approach requires
choosing a time frame of integration (see section
2.7 on metrics for more discussion of time frames).
Third, there are increased uncertainties because the
results depend on decisions regarding background
concentrations and are more sensitive to model
factors such as atmospheric chemistry and carbon
cycles. In contrast, analyses based on existing
concentrations are less uncertain because for the
most part they depend on measured concentrations.
In general, this is more important for the long-
lived GHGs than for short-lived substances such
as BC, and thus becomes an issue for purposes of
comparisons.
The second issue in addressing climate impacts
is whether to consider radiative forcing as an
endpoint, or whether to actually calculate impacts
on temperatures. The advantage to examining
temperatures is that they are more immediately
relevant to human experience than the more
abstract "radiative forcing". Also, though radiative
forcing is approximately additive in most cases,
there are some exceptions: as discussed in section
2.6.1.4, temperatures are actually much more
sensitive to snow albedo forcing than to other
forcings.
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Black Carbon and Its Effects on Climate
However, there are two disadvantages to the
temperature approach. The first is that, again,
fewer studies use this approach, in part because
the models used to calculate temperature are more
computationally intensive than those that calculate
radiative forcing. The second is that the relationship
between radiative forcing and temperature, while
generally linear, is very model dependent. Every
model has an inherent "climate sensitivity", and
therefore when comparing the impacts of BC from
one model to the next, differences in this climate
sensitivity can make it difficult to understand if
differences between the models are due to BC
physics or to the general model response to any
forcing.
2.6.1 Global and Regional Radiative Forcing
Effects of BC: Overview
This section provides an overview of the different
effects of BC on radiative forcing based on the best
estimates in the current literature. These effects
include direct forcing (direct absorption of solar
or terrestrial radiation), snow/ice albedo forcing
(forcing that results from the darkening of snow and
ice), and indirect forcing (a range of forcing effects
resulting from impacts on clouds, including changes
in cloud lifetime, reflectivity, and composition).
Section 2.6.1.1 describes the overall net effects
of BC on radiative forcing, when the effects of
these different types of forcing are accounted for.
The subsequent sections present more detailed
information regarding the state of knowledge
regarding the specific effects of BC on direct forcing,
snow/ice albedo forcing, and indirect forcing at the
global and regional scales. Each section provides
a summary of the findings of recent studies,
explanations of the differences among estimates,
and characterizations of key remaining uncertainties.
In evaluating estimates of the effects of BC on
radiative forcing, it is necessary to consider several
caveats. For example, it is important to differentiate
among estimates with respect to the baseline time
period used to define the radiative forcing estimates.
The radiative forcing estimates (and other climate
effects) are often expressed as a comparison to
a given historical level rather than with respect
to present day or in terms of the anthropogenic
influence compared to total forcing. However, these
assumptions are not always stated clearly. Also, it is
important to differentiate with respect to the types
of BC emissions included. For example, many studies
exclude open biomass burning. The inclusion or
exclusion of BC from wildfires and other sources of
open biomass burning will affect the estimates of
net BC effects. In addition, because some studies
evaluate the climate effects of BC as it co-occurs with
other aerosol chemical species, such as OC, sulfates
and nitrates, while others do not, it is important
to distinguish studies where BC is estimated
individually from studies where BC is estimated as
part of an aerosol mixture. In the following sections,
it is indicated whether the radiative forcing estimates
include co-occurring OC and other species and how
these other pollutants influence estimates of BC's
global and regional climate impacts, when possible.
2.6.1.1 Net Forcing
As is described in more detail in the following
sections, the different kinds of forcing involve
different mechanisms of action and can have
offsetting climate effects. For example, direct
effects are associated with positive forcing, while
most (but not all) indirect effects are thought to
result in negative radiative forcing. This section
provides an overview of estimates of the direction
and magnitude of the net effect of BC on radiative
forcing when the direct forcing, snow/ice albedo
forcing, and indirect forcing effects of BC are
summed, and identifies key factors that contribute
to variability in these estimates
There is a range of quantitative estimates in the
literature for global average radiative forcing due
to BC. Most studies indicate that due to the direct
and snow/ice albedo effects, the net effect of BC on
climate is likely to be warming. However, because
of the large remaining uncertainties regarding
interactions of BC with clouds, it is difficult to
establish quantitative bounds for estimating global
net impacts of BC, or even to completely rule out the
possibility of a net negative effect.
The most widely utilized estimates of forcings for
GHGs come from the IPCC's Fourth Assessment
Report, which was issued in 2007. The IPCC also
estimated forcing due to BC based on a review of
the scientific studies available at the time, though
estimates from more recent studies surveyed in
this EPA report differ somewhat as detailed further
below. The IPCC estimated a direct radiative forcing
of +0.34 W m 2for BC, making BC third only to
CO2 and CH4. In addition, the IPCC estimated BC's
snow/ice albedo forcing to be +0.1 W m 2 (see
Figure 2-10). Other aerosols were generally shown
to have a cooling influence on climate. The IPCC
estimates of negative direct radiative forcing due
to OC and sulfates are also shown in Figure 2-10.
Indirect effects for all aerosols, including BC, are
also estimated to result in net negative forcing due
to increased reflectivity of clouds ("cloud albedo
effect"). The IPCC did not provide quantitative
estimates of the effect of aerosols on other
properties of clouds (such as lifetime, stability, etc.),
Report to Congress on Black Carbon
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Chapter 2
Long-Lived
Greenhouse
Gases
Aerosols
and
Precursors
CO,
CH,
N2O
Black Carbon
Organic Carbon
SO,
Aerosols
Black Carbon (Direct)
Black Carbon (Snow/Ice Albedo)
Organic Carbon I
i i
Sulfate (Direct)
Cloud Albedo Effect
i i
j i
-1 -0.75 -0.5 -0.25 0 0.25 0.5 0.75
Estimated Forcing (W/m2)
1.25
1.5
Figure 2-10. Components of Global Average Radiative Forcing for Emissions of Principal Gases,
Aerosols, and Aerosol Precursors, based on IPCC estimates. Values represent global average radiative
forcing in 2005 due to emissions and changes since 1750. Total radiative forcing for CH4 includes the effects
of historical CH4 emissions on levels of tropospheric O3 and stratospheric H2O, and the CO2 oxidation
product of CH4from fossil sources. Similarly, total radiative forcing for N2O includes the effect of historical
N2O emissions on levels of stratospheric O3. The IPCC does not report an overall uncertainty for the net
contribution to forcing of individual GHG emissions. However, based on the uncertainties provided for
the individual components of these contributions, the uncertainty in forcing from CO2 and N2O emissions
is extrapolated as being approximately 10% and approximately 20% from CH4emissions. Uncertainty in
direct forcing is ±0.25 W nv2 for BCand ±0.20 W nr2 for both OCand SO2. The range of forcing for the cloud
albedo effect is -1.8 to -0.3 W nr2. (Adapted from Figure 2.21 of Forster et al., 2007)
or quantify the indirect effects of individual aerosol
species (such as BC) separately. As a result, there is
substantial uncertainty in the IPCC's estimates of net
forcing for BC.
In addition to the estimates compiled by the
IPCC (2007), many other studies have attempted
to estimate the global average radiative forcing
attributable to BC. An examination of the results of
these studies, as summarized in Figure 2-11, indicates
that the direct effect and the snow/ice albedo effect
of BC are positive, though the magnitude of these
effects is uncertain. The figure shows the range of
central estimates from the included studies (solid
box) as well as the highest and lowest uncertainty
estimates from those studies (error bars) for both
the direct effect and the snow and ice albedo effect.
As discussed further below, a number of studies
have estimated BC's direct radiative forcing to be
higher than the IPCC estimate (Sato et al., 2003;
Ramanathan and Carmichael, 2008).
The biggest source of uncertainty about the net
forcing effect of BC is the magnitude of the cloud
effects of BC. (Cloud effects are discussed in detail
in section 2.6.1.3.) The limited number of studies in
the literature allow for statements on the direction
(e.g., warming or cooling) of some of these forcings,
but not their magnitude, as shown in Figure 2-11.
The impact on cloud lifetime and albedo is likely
cooling. The interactions with mixed-phase and ice
clouds are likely to be warming. Semi-direct effects
are uncertain, and existing studies differ on the
definition and net influence of the effect. The cloud
absorption effect is positive.
In light of the large remaining uncertainties about
the magnitude (and in some cases the sign) of the
different forcing effects of BC, particularly with
regard to the cloud interaction effects, this EPA
report does not assign a range to the magnitude of
the net effect beyond noting that it is very likely to
be positive (however, a net negative effect cannot
be excluded). As indicated in Figure 2-11, the
34
Report to Congress on Black Carbon
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Black Carbon and Its Effects on Climate
O)
o
LL
-5
CD
Q.
P
0)
O
"o
0
Q.
£
Direct
Snow and
Ice Albedo
-1.
Cloud Lifetime
and Albedo
Mixed Phase
and Ice
Semi-Direct
i i i
i i i
i i i
i i i
i i i
i i i
i i i
25 -1 -0.75 -0.5 -0
9
| ^ |
i i i i i
i • i
hi — I
i i i i i
i i i i i
25 0 0.25 0.5 0.75 1 1.
9
9
25
y
o
cu
Q.
Net
0 +
Estimated Forcing (W/m2)
Figure 2-11. Estimates of Radiative Forcing from BC Emissions Only. The boxes indicate ranges of central
estimates from the papers identified in this report, with error bars indicating the highest and lowest
uncertainty estimates from those papers. Estimates are based on a synthesis of results from eleven studies
that considered the direct forcing effects of BC emissions and six studies that considered reduction in snow
and ice albedo from BC emissions. The range for the snow and ice albedo bar does not include the effects
of the higher efficacy of the snow albedo effect on temperature change (forcing efficacy is described in
section 2.6.1.4). The studies of indirect and semi-direct radiative forcing effects due to BC emissions are
not sufficiently comparable in scope and approach to combine the estimates. As a result, only the likely
direction of forcing is presented. Two other warming effects, the cloud absorption effect and the water
vapor effect, have recently been discussed in the literature but are not included here. (Source: U.S. EPA)
estimates of direct and snow/ice albedo forcing are
likely to be positive, but additional work is needed to
determine the extent to which these positive forcing
effects are offset by indirect effects, semi-direct
effects, and other effects on clouds.
The recent UNEP/WMO assessment (UNEP and
WMO, 2011a) evaluated a number of recent studies
to investigate the net effect of BC on climate. Based
on estimates of BC forcing due to direct, indirect,
and snow/ice albedo effects, UNEP/WMO estimated
that the global average net forcing was likely to be
positive and in the range of 0 to 1 W rrr2, with a
central value of 0.6 W rrr2 (this estimate included an
enhanced efficacy factor for the snow/ice albedo
effect: forcing efficacy is described in section
2.6.1.4). The UNEP/WMO assessment estimates
for the different forcing effects of BC are shown in
Table 2-3. In selecting a central estimate for net BC
forcing, the authors of the UNEP/WMO assessment
noted the very strong negative forcing due to
total aerosols,9and the current lack of quantitative
9The IPCC estimated the net forcing due to all aerosols as
-1.2 W rrr2 (including a direct forcing effect of -0.5 W m 2
(-0.9 to -0.10 W m 2) and a cloud albedo effect of -0.7 W m 2
(-1.8 to -0.3 W m 2) (IPCC, 2007). However, these estimates
are assigned a low level of scientific understanding by IPCC.
More recent estimates in the literature continue to span
a large range: Bauer and Menon (2012) estimate the net
aerosol effect as -0.6 W m 2 (including direct -0.5 W m 2 and
indirect -0.1 W m2), while Hansen et al. (2011) argue the
effect is more strongly negative (-1.6 W rrr2 ±0.3)
Report to Congress on Black Carbon
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Chapter 2
Table 2-3. Summary of UNEP/WMO Assessment Estimates of Radiative Forcing Effects of BC. (Adapted
from UNEP and WMO, 201 la, Table 3-1)
Forcing Range (W/m2) Central Values (W/m2) Mean Value (W/m2)
Direct
Semi-direct and indirect
Deposition
Total BC
0.2 to 0.9
-0.6 to 0.5
(0.01 to 0.08)
corrected for efficacy (x5)
0.05 to 0.4
-0.3 to 1.8
0.3 to 0.6
-0.4 to 0.4
(0.01 to 0.05)
0.05 to 0.25
0.0 to 1.3
0.0 to 1.0 with constraints on total
aerosol forcing
0.45
0.0
0.03
0.15
0.5
0.6 including efficacy of deposition
estimates for all processes. In light of these
constraints, they argued that net effective BC forcing
is unlikely to exceed 1 W rrr2.
There remains a strong need for further research
to provide better estimates of the forcing effects of
BC and to reduce remaining uncertainties. Work in
progress by a consortium of researchers under IGAC/
SPARC (currently being prepared for submission
to an academic journal with expected publication
later in 2012) will likely provide more definitive
quantitative bounds on the BC cloud interaction
effects, and the net effects overall.
2.6.1.1.1 Factors that Contribute to Variability in
Estimates
There are a number of factors that may contribute to
the lack of consensus among modeled estimates of
net global average radiative forcing from BC. Koch
et al. (2009) attributed the range of estimates to
differences in the aerosol microphysical calculations
in the models (i.e., different estimates of how much
solar radiation each unit of BC absorbs). The authors
also pointed out key differences in models, such as
the assumed values of various physical properties,
and differences in the representation of vertical
transport and cloud effects.
Variability in the estimates may also arise due to
differences in experimental design and how the
values are reported. Radiative forcing is commonly
measured and reported as top-of-the-atmosphere
(TOA) radiative forcing which captures all variations
in energy over the entire atmosphere. This is
appropriate for the well-mixed, long-lived GHGs,
but perhaps not for BC, which exhibits high spatial
variability. For example, the vertical distribution of
BC in the atmospheric column and interactions with
clouds lead to inputs of energy at different altitudes
compared to the input of energy due to GHGs (see
Ramanathan et al., 2001, and references therein).
Climate effects are also sensitive to the location of
the BC emissions. For example, Arctic sea ice melting
may be accelerated by BC emissions from northern
latitudes, as discussed later in this chapter. Finally,
radiative forcing metrics that focus on specific
species do not generally capture co-pollutant
interactions, which are very important for BC.
2.6.1.1.2 Regional Dynamics
Studies focusing on global average radiative forcing
may overlook key regional dynamics associated
with BC as a spatially heterogeneous pollutant.
Many studies have found that BC's regional
climate impacts are more pronounced than the
contributions of BC to global average temperature
change. In addition, certain regions of the world are
more sensitive to or more likely to be affected by BC
forcing, either due to transport and deposition (e.g.,
the Arctic) or high levels of aerosol pollution in the
region (e.g., Asia). Global average radiative forcings
for BC hide much of the regional variability in the
concentrations and impacts. Note, however, that
regional variability of BC forcing may exaggerate
the regional variability of impacts, as temperature
impacts usually occur over a larger area and longer
time period than the forcing effects.
2.6.1.2 Direct Forcing
The direct effect of BC is to absorb solar radiation.
As mentioned above and as shown in Figure 2-12,
the IPCC (2007) estimated the global average
radiative forcing of BC from all sources at +0.34
(±0.25) W m2. A subset of this forcing due to BC
from fossil fuel combustion (mainly coal, petroleum
and gas fuels) was estimated to be +0.2 (±0.15) W
m2. Most studies published since the IPCC report
have reported higher direct forcing values. The
UNEP/WMO assessment (UNEP and WMO, 2011a)
estimated that direct forcing of BC is most likely to
be within the range 0.3 to 0.6 W rrr2. Additional work
36
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Black Carbon and Its Effects on Climate
Sources of Emissions
Name of Study
All Sources
(BB, BF, FF)
FFand BF
BB only
FFonly
Bond etal. (2011)
Ramanathan (2010)
Ramanathan and Carmichael (2008)
Forster et al. (2007) from IPCC (2007)
Chung and Seinfeld (2005)
Sato et al. (2003)
Jacobson(2001)
Bond etal. (2011)
Myhre et al. (2009)
Reddy and Boucher (2007)
Bond etal. (2011)
Hansen et al. (2005)
Forster et al. (2007) from IPCC (2007)
Hansen et al. (2005)
Figure 2-12. Estimates of
Direct Radiative Forcing
from BC Emissions Only.
These values represent the
range of estimates in the
peer reviewed literature;
however, they are not all
directly comparable. Some
are based on different
estimates of BC emissions,
include different sectors,
and present the forcing with
respect to different baseline
time periods (e.g., 1750, post-
industrial, present day). Note:
BB = open biomass burning;
BF = biofuels; and FF = fossil
fuels. (Source: U.S. EPA)
0.25 0.5 0.75 1
Estimated Forcing (W/m2)
1.25
is underway to try to develop a new central estimate
for these direct impacts (IGAC/SPARC, forthcoming).
Assumptions about mixing state (e.g., internal/
external) are critical to the results. As noted in
section 2.5, studies that have incorporated internal
mixing into the calculations of direct radiative
forcing for BC yield higher forcing than those that
do not, and these models are considered to be more
realistic. Simulations by Jacobson (2001) found that
accounting for internal mixing of BC in aerosols
increases the estimated absorption and warming by
BC by a factor of two. Koch et al. (2009) accounted
for this underestimation of absorption by BC in
older models by doubling the ensemble average
from a 17-model intercomparison project (Schulz
et al., 2006), resulting in a global average BC direct
radiative forcing of roughly +0.5 W rrr2. Bond et al.
(2011) combined forcing results from 12 models
to use the best estimates for mixing and transport
in those models. Based on this analysis, and using
the same emissions estimates used by the models
assessed in the IPCC reports, Bond et al. (2011)
found a total forcing of +0.40 W rrr2, or 18% higher
than the IPCC estimate, which they attributed to the
fact that the IPCC estimate includes some models
that do not include enhanced absorption due to
internal mixing. Note that Bond et al. differentiate
"anthropogenic emissions" (post-1750, including
open burning) and total BC emissions, calculating
the total forcing from the latter to be +0.47 W rrr2.10
In addition to mixing state, other factors that can
lead to variation in modeled radiative forcing effects
of BC include particle size and removal (e.g., by
precipitation). Vignati et al. (2010) describe how a
modeled reduction in wet removal of BC from the
atmosphere by 30% results in a 10% increase in
BC's atmospheric lifetime. In some cases, mainly in
work based on observational constraints from the
Aerosol Robotic Network (AERONET) ground-based
sunphotometer network, much higher values have
been reported. Sato et al. (2003) inferred a forcing
of 1 W m2 based on these observational constraints.
Chung et al. (2005) and Ramanathan and Carmichael
(2008) combined the AERONET results with satellite
data and report an estimated global average
radiative forcing for BC of +0.9 W rrr2, with a range
of +0.4 to +1.2 W m^2. While most recent studies find
global forcing higher than the IPCC, a discrepancy
remains between the very high observationally
constrained results and model results (even those
10 Ramanathan (2010) suggests that Bond et al. (2011)
underestimate the total forcing, and suggests a central
value of +0.55 with a 90% confidence interval of +0.2 to
+0.9 W m-2.
Report to Congress on Black Carbon
37
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Chapter 2
that include internal mixing and therefore produce
higher values). Bond et al. (2011) hypothesized that
the higher forcing in the observationally constrained
results could result from higher emissions than in the
model work. The exact cause of these differences,
however, has not been isolated.
Compared to global radiative forcing, fewer studies
have reported regional direct radiative forcing by
BC. Studies such as Bond et al. (2011) show the
geographic distribution of direct forcing from all
sources of BC emissions (Figure 2-13). They found
the largest forcing over South and East Asia and
parts of Africa. Myhre et al. (2009), who considered
only fossil fuel and biofuel BC, also found the
largest forcing over South and East Asia. Other
work such as Chung and Seinfeld (2005) showed
similar patterns with higher forcing in Central and
South America (the Amazon basin) and sub-Saharan
Africa due to the inclusion of biomass burning
emissions. Chung and Seinfeld (2005) report a range
of +0.52 to +0.93 W rrv2 for externally and internally
mixed BC respectively, averaged over the Northern
Hemisphere. Their earlier work also suggests a
strong seasonal cycle which peaks in May at +1.4
W m2 (Chung and Seinfeld, 2002). For the Southern
Hemisphere, Chung and Seinfeld (2005) estimate a
range of +0.15 to +0.23 W m2. Reddy and Boucher
(2007) calculated the influence of regional BC
emissions on the global average radiative forcing.
The largest contribution to global TOA BC radiative
forcing came from East Asia (+0.08 W m 2). The
global average forcing due to North American BC
emissions in this study was +0.02 W m2.
2.6.1.3 Cloud-related Forcings
The net effect of particles on climate via impacts
on clouds is highly uncertain (IPCC, 2007). There
are several different kinds of cloud effects that are
important for radiative purposes, as summarized in
Table 2-4. These cloud effects contribute to changes
in the radiative balance of the atmosphere, and also
influence climatic factors such as precipitation and
dimming (section 2.6.3).
Since cloud droplets are formed when water vapor
condenses onto a particle, many types of particles
can affect the formation and microphysics of clouds.
Emissions of aerosols into the atmosphere increases
the number of particles on which cloud droplets
can form, resulting in more and smaller cloud
droplets. These additional cloud droplets produce
brighter, more reflective clouds (Twomey, 1977). This
generally results in surface cooling by preventing
sunlight from reaching below the cloud to the
Earth's surface (see also section 2.6.3.1 on surface
dimming). This increase in reflectivity of the clouds
has been termed the "the first indirect effect" or
the "cloud albedo effect". In addition, the smaller
Figure 2-13. Direct Radiative Forcing (W m2) of BC from All Sources, simulated with the Community
Atmosphere Model. (Bond et al., 2011)
38
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Black Carbon and Its Effects on Climate
Table 2-4. Overview of the Different Aerosol Cloud Effects. This summary applies to all aerosols and is
not BC-specific, though BC is expected to participate to some extent in all of these mechanisms. Net effect
refers to top-of-the-atmosphere radiative forcing. Scientific understanding is based on IPCC terminology.
(Adapted from IPCC, Denman et al., 2007, Table 7.10a)
Effect al«dYT Process Net Effect J?0*"1'*1 ,, Jf^l
Affected Magnitude Understanding
Cloud albedo effect
(first indirect effect)
Cloud lifetime effect
(second indirect
effect)
Semi-direct effect
Glaciation indirect
effect
Thermodynamic
effect
All clouds
All clouds
All clouds
Mixed-phase
clouds
Mixed-phase
clouds
Smaller cloud particles reflect
more solar radiation
Smaller cloud particles decrease
precipitation so increase lifetime
Absorption of solar radiation
by absorbing aerosols changes
atmospheric stability and cloud
formation
An increase in ice nuclei due
to some aerosols increases
precipitation
Smaller cloud droplets delay
freezing causing super-cooled
clouds to extend to colder
temperatures
Cooling
Cooling
Cooling or
warming
Warming
Cooling or
warming
Medium
Medium
Small
Medium
Medium
Low
Very low
Very low
Very low
Very low
cloud droplets are less likely to aggregate sufficiently
to form rain drops, which changes precipitation
patterns and increases cloud lifetime (Albrecht, 1989)
(see also section 2.6.3.2 on precipitation impacts).
This has been called the "second indirect effect" or
the "cloud lifetime effect". In general, the cloud
albedo effect and the cloud lifetime effect are
estimated to lead to cooling. However, Ramanathan
(2010) asserts that the empirical evidence shows a
positive forcing (warming) over land regions.
The magnitude and sign of the radiative effects
depend on whether the clouds are composed
of liquid droplets, ice particles, or a mix of ice
and liquid droplets, and on the composition of
the aerosol particles. In certain kinds of "mixed-
phase clouds" (clouds with both ice and water),
smaller droplets cause a delay in the freezing of
the droplets, changing the characteristics of the
cloud; however, the IPCC was not able to determine
whether this "thermodynamic effect" would result
in overall warming or cooling (Denman et al., 2007).
The "semi-direct effect" is specific to BC and other
absorbing aerosols, while the "glaciation indirect
effect" appears to be important for aerosols, such
as BC and mineral dust, which are not especially
hydrophilic. The semi-direct effect refers to the
heating of the troposphere by absorbing aerosols,
affecting the relative humidity and stability of the
troposphere, which in turn affects cloud formation
and lifetime (IPCC, 2007; Ackerman et al., 2000).
Older literature refers to the semi-direct effect as
cloud burn-off (i.e., a decrease in cloud formation)
from BC within the cloud layer. The definition was
extended to include all effects on cloud formation
and lifetime as other studies have found that
humidity and stability effects from BC above
and below clouds can cause both increases and
decreases in clouds (Koch and Del Genio, 2010). The
IPCC did not assign a sign to the net forcing of the
semi-direct effect (Denman et al., 2007).
More recently, Koch and Del Genio (2010) found
in their review of the literature that most model
studies generally indicate a global net negative
semi-direct effect (i.e., the effect of atmospheric
heating by absorbing aerosols on cloud formation
and lifetime causes net cooling). This was observed
despite regional variation in the cloud response
to absorbing aerosols (such as BC), and resulting
regional differences in warming and cooling from
the semi-direct effect. In contrast to Koch and Del
Genio, Jacobson (2010) found that the semi-direct
effect is positive. This difference may be due in part
to the more inclusive definition used by Koch and
Del Genio. Isaksen et al. (2009) reported a range of
-0.25 to +0.50 W m2.
While the sign of the semi-direct effect is therefore
in question, the glaciation effect is very likely a
warming effect, though it occurs only in some
mixed-phase clouds. This indirect effect is caused
by BC aerosols (and some other particles such
Report to Congress on Black Carbon
39
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Chapter 2
as mineral dust) serving as ice nuclei in a super-
cooled liquid water cloud, thereby enabling
precipitation rather than delaying it (Denman et al.,
2007; Lohmann and Hoose, 2009). However, some
preliminary work (e.g., Penner et al., 2009) suggests
that the effect of soot on ice nuclei and their
subsequent effect on cirrus clouds could offset some
of the warming resulting from the glaciation effect.
Most estimates of the forcing from aerosol indirect
effects are based on all aerosol species (e.g., total
PM) and are not estimated for individual species
(e.g., BC alone). The net indirect effect of all aerosols
is estimated as a negative value. The IPCC (Forster et
al., 2007) estimated the change in cloud albedo due
to all aerosols to have a radiative forcing of -0.7 W
rrr2, with a 5 to 95% confidence range of -0.3 to -1.8
W m"2 and a low level of scientific understanding.11
A recent study by Bauer and Menon (2012) found a
smaller net cloud interaction effect due to aerosols
of only -0.1 W rrr2.12
There are some BC-specific effects that do not fall
into the IPCC categories from Table 2-4, which have
been identified in a limited number of studies. The
first is called the "cloud absorption effect." This effect
is the result of BC particles being included between
and within cloud particles, increasing absorption (or
decreasing reflectivity) of the clouds. Jacobson (2010)
suggests that the inclusion of this cloud absorption
effect may increase warming from BC by as much as
75%. Additionally, recent experimental work above
the Amazon indicates that cloud albedo will increase
with increasing aerosol loading (as predicted by
the cloud albedo effect) only as long as the aerosol
loading is smaller than a critical threshold. Above
this threshold, additional aerosols from biomass
burning actually lead to a decrease in cloud albedo,
attributed to inhibition of cloud formation by
absorbing aerosols (Ten Hoeve et al., 2011; Koren
et al., 2008). The second effect that does not fall
into the more common IPCC categories is called the
"BC-water vapor effect" (Jacobson, 2006, 2010). This
effect is based on an increase in water vapor due to
inhibition of precipitation from clouds, related to the
cloud absorption effect. However, BC can also lead
to a decrease in relative humidity in some regions
where the increase in atmospheric temperature at
altitude does not lead to increased evaporation at
11 The IPCC definition of "level of understanding" is a qualitative
measure based on a combination of the quantity of evidence
available and the degree of consensus in the literature.
12 Bauer and Menon estimated the indirect cloud albedo effect as
-0.17 W m-2 and the semidirect effect as -0.10 W m-2, but noted
these effects "can be isolated on a regional scale, and they often
have opposing forcing effects, leading to overall small forcing
effects on a global scale."
the surface. It is unclear what the magnitude of this
effect is.
To summarize, it is unclear to what extent BC
contributes to the overall aerosol indirect effect.
As a result, this report does not assign any central
estimate or even a range of possible values for the
role of BC in the overall indirect aerosol effect. BC's
role in the first and second indirect effects (cloud
albedo and cloud lifetime effects) is likely to be
cooling, but possibly to a lesser extent than for
other aerosols. Although freshly emitted, externally
mixed BC particles are hydrophobic and would be
less active cloud condensation nuclei (CCN), aging
may increase their ability to serve as CCN (Dusek
et al., 2006). Recent work (e.g., Bauer et al., 2010)
using models with a more explicit representation of
aerosol mixing than older models suggests that the
role of BC in the indirect effect may be greater than
previously thought. Similarly, BC may also participate
in the thermodynamic indirect effect for mixed-
phase clouds, but whether this effect is net warming
or cooling is still uncertain. BC has a primary role in
the semi-direct effect, but this effect may produce
warming or cooling depending on conditions.
Finally, BC particles may contribute to warming
from the glaciation indirect effect in mixed-phase
clouds, the cloud absorption effect, and the water
vapor effect, but the magnitude of these effects are
uncertain. A comprehensive, quantitative estimate
of the net effect of BC would require an assessment
of the likely bounds of these cloud effects. While
this EPA report does not assess those bounds, the
UNEP/WMO assessment (UNEP and WMO, 2011a)
found that the most likely case was that the cloud
interaction effects resulting from BC alone would
cancel out (mean value of 0.0 W rrv2), but with a
central range of -0.4 to +0.4 W rrr2. The forthcoming
IGAC/SPARC study will present a more definitive
bound on these effect estimates. Studies are just
beginning to estimate indirect or semi-direct
radiative effects for BC at a regional level (e.g., Bauer
and Menon, 2012).
2.6.1.4 Snow and Ice Albedo Forcing
BC deposited on snow and ice leads to positive
radiative forcing. It darkens the surface which
decreases the surface albedo, and it absorbs
sunlight, heating the snow and ice (Warren and
Wiscombe, 1980). The snow and ice albedo effect
is strongest in the spring because snow cover is
at its greatest extent, and spring is a season with
increased exposure to sunlight (Flanner et al., 2009).
BrC has also been found to contribute to snow and
ice albedo forcing (Doherty et al., 2010). Chapter 5
also addresses observations of BC in snow in more
depth.
40
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Black Carbon and Its Effects on Climate
There are a number of estimates of the magnitude
of radiative forcing due to the snow albedo effect
(see Figure 2-14). In a modeling study, Hansen and
Nazarenko (2004) estimated the global average
radiative forcing of BC on snow and ice to be +0.16
W m2 for what they considered to be the most
realistic of the four cases that were simulated in their
study. In later work, Hansen et al. (2007b) lowered
this estimate to +0.05 W rrr2, with a probable range
of 0 to +0.1 W m2 (Hansen et al., 2007b). Relying on
these studies, the IPCC (Forster et al., 2007) adopted
a best estimate for the global average radiative
forcing of deposited BC on snow and ice of +0.10
(±0.10) W m^2, although the authors acknowledged
a low level of scientific understanding regarding
this effect. In more recent work, Flanner et al. (2007)
estimated the average forcing of BC on snow and ice
(from fossil fuels and biofuels) at +0.043 W rrr2, of
which +0.033 W rrr2 was attributed to BC from fossil
fuels. When biomass burning was included in the
calculation, the forcing of BC on snow and ice was
estimated to be approximately +0.05 W m2. Bond et
al. (2011) estimated a global forcing of +0.047 W rrv2,
of which 20% was calculated to occur in the Arctic
(defined as north of 60 degrees), and suggested that
more mechanistic studies in general yield estimates
lower than the central IPCC estimate of +0.1 W m2. In
line with these more recent studies, the UNEP/WMO
assessment (UNEP and WMO, 2011a) estimated 0.05
W rrr2, with a range of 0.01 to 0.10 W rrr2, for the
snow and ice albedo effect.
Hansen et al. (2007b) also investigated the
"effectiveness" (or "efficacy") of the snow albedo
forcing. This is a relative measure of positive
feedback effects that occur with BC, compared to
the feedbacks that occur with warming due to CO2
forcing. They calculated that the radiative forcing
from decreases in surface albedo is 2.7 times more
effective at warming than radiative forcing from CO2.
This is a result of the energy absorption from the BC
being directly applied to melting snow rather than
spread throughout the height of the atmosphere.
BC particles left behind in melting surface snow can
concentrate and further reduce the surface albedo
(see section 5.6). Furthermore, BC deposited on ice
and snow will continue to have radiative effects
as long as the BC remains exposed (i.e., until the
snow melts away or fresh snow falls). Melting snow
can expose a dark surface, leading to a positive
feedback. Flanner et al. (2007) found a larger
efficacy of 3.2, with an uncertainty range of 2.1 to
4.5. Based on the more recent work of Flanner (2009)
and Koch (2009), the UNEP/WMO assessment (UNEP
and WMO, 2011a) adopted an even larger effective
forcing of a factor of five. Flanner et al. (2011) also
found that observed Northern Hemisphere snow
retreat between 1979 and 2008 (from all causes)
Sources of Emissions
Study
All Sources
(BB, BF, FF)
FF and BF
FF only
Bondetal. (2011)
Ramanathan (2010)
Hansen et al. (2007)
Flanner et al. (2007)
Forster et al. (2007) from IPCC (2007)
Bondetal. (2011)
Flanner etal. (2007)
Flanner et al. (2007)
0 0.05 0.10 0.15 0.20 0.25
Estimated Forcing (W/m2)
0.30
Figure 2-14. Estimates of Snow and Ice Albedo Radiative Forcing Effects from BC Emissions Only. Note:
BB = open biomass burning; BF = biofuels; and FF = fossil fuels. Dashed lines indicate the estimated range
of snow and ice albedo radiative forcing when forcing efficacy is considered. Hansen et al. (2005) estimate
an efficacy of 170% for the snow and ice albedo radiative forcing effect of BC emissions. (Source: U.S. EPA)
Report to Congress on Black Carbon
41
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Chapter 2
90
45
0
-45
-90
0.025
Figure 2-15. Snow and Ice Albedo Forcing by BC, simulated with the Community Atmosphere Model.
(Bond etal., 2011)
would be consistent with a total albedo feedback on
the order of +0.45 W rrr2. This suggests that albedo
feedback is a larger process than represented in
most climate models.
For snow and ice, there is evidence that all
atmospheric PM, including all mixtures of BC
and OC, increases the net solar heating of the
atmosphere-snow column (Flanner et al., 2009).
This means that mixtures of BC and OC that are
transported over snow-covered areas may have a net
warming influence regardless of the ratio of the two
compounds (although Flanner et al. did not include
cloud effects). This is in contrast to direct radiative
forcing estimates, which are strongly influenced by
the ratio of BC to other cooling PM components such
as OC. Flanner et al. (2009) also found that fossil-
fuel and biofuel BC and OM emissions contributed
almost as much to springtime snow loss in Eurasia as
did anthropogenic CO2. The size and composition of
the deposited particles affects how long they remain
on or near the surface where they are able to reduce
albedo.
Snow and ice albedo forcing is confined to areas
with snow and ice cover (approximately 7.5%-15%
of Earth's surface; see section 5.6). Thus, global
average forcing estimates do not convey the
significant spatial and temporal variability in the
radiative forcing of BC on snow and ice. Radiative
forcing from changes in snow and ice albedo from
BC are estimated to be much larger than the global
averages for much of Northern and Eastern Europe,
Russia, and China. These effects are especially
pronounced in the Arctic and the Himalayas. Flanner
et al. (2007) calculated an average forcing of BC on
snow and ice of +1.5 W rrr2 in the Tibetan plateau,
with instantaneous forcings13 of up to +20 W rrv2 in
the spring. These high values are due to the large
amount of mountain snow and ice cover as well as
the proximity to high emissions of BC from parts of
China and the Indian subcontinent. Large radiative
forcing values have also been estimated over the
Arctic. Hansen and Nazarenko (2004) calculated an
average forcing due to BC on snow and ice of +1 W
m2 in the Arctic compared to +0.3 W m2 over the
Northern Hemisphere as a whole. However, these
estimates are based on global numbers that were
reduced by a factor of three in later papers (Hansen
et al., 2007b). The full spatial distribution of forcing
by BC on snow and ice as simulated by Bond et al.
(2011) is shown in Figure 2-15. The effects of BC on
the Arctic and the Himalayas are described in more
detail in sections 2.6.4 and 2.6.5, below.
2.6.1.5 The Radiative Forcing Effects of OC and
other Co-Pollutants
Although BC is mixed with other pollutants, both
at the point of emission and in the atmosphere,
most studies examine the impact of different types
13 Instantaneous radiative forcing refers to the flux at the
tropopause, rather than forcing averaged over a longer time period.
42
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Sources of Emissions
Black Carbon and Its Effects on Climate
Study
Bond et al. (2011)
Forster et al. (2007) from IPCC (2007)
Chung and Seinfeld (2002)
Bond etal. (2011)
Myhre et al. (2009)
Hansen et al. (2005)
Bond etal. (2011)
Hansen etal. (2005)
Forster et al. (2007) from IPCC (2007)
Hansenetal. (2005)
-0.40 -0.30 -0.20 -0.10 0
Estimated Forcing (W/m2)
0.10
Figure 2-16. Estimates of Direct Radiative Forcing from OC Emissions Only. Note: BB = open biomass
burning; BF = biofuels; and FF = fossil fuels. (Source: U.S. EPA)
Sources of Emissions
Study
All Sources
(BB, BF, FF)
FF and BB
BB only
FF only
i
^^*
L
F
HH
, , ,
r~> I
I
.
^HjH |
*-\
Forster et al. (2007) from IPCC (2007)
Hansen et al. (2005)
Myhre et al. (2009)
Forster et al. (2007) from IPCC (2007)
Hansen et al. (2005)
Hansen et al. (2005)
Haywood and Shine (1995)
-1 -0.75 -0.5 -0.25 0 0.25 0.5 0.75 1
Estimated Forcing (W/m2)
Figure 2-17. Estimates of Direct Radiative Forcing from BC and OC Emissions. Note: BB = open
biomass burning; BF = biofuels; and FF = fossil fuels. Forster et al. (2007) - from IPCC (2007) - estimate
the uncertainty surrounding estimates of direct radiative forcing from BCandOC independently. For this
reason, the uncertainty surrounding the combined estimated direct radiative forcing from BC and OC
emissions from all sources according to Forster et al. (2007) is omitted. (Source: U.S. EPA)
Report to Congress on Black Carbon
43
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Chapter 2
of aerosols in isolation. Only a limited number of
studies consider the impacts of co-pollutants, and
most of these studies have focused on OC rather
than all aerosol species. Figure 2-16 shows estimates
of direct radiative forcing for OC from a number of
studies. As indicated in Figure 2-16, OC emissions
from all sources are estimated to have net cooling
impacts. For example, the IPCC (2007) estimated the
negative direct radiative forcing of OC aerosols from
all sources at -0.19 (±0.20) W rrr2 and from fossil fuel
alone at -0.05 (±0.05) W m 2.
When OC and BC emissions are combined, the
estimates of global average direct radiative forcing
are generally positive. Figure 2-17 shows estimates
for BC and OC combined from different sources.
Here, the total direct radiative forcing from BC and
OC emissions from all sources was estimated by
IPCC (2007) at approximately +0.15 W m2 (global
average), and even biomass burning aerosols were
estimated to have a positive net forcing of +0.03
(±0.12) W m"2. Another study that calculated a net
forcing from BC and OC from all sources reported
a net global average forcing of about +0.27 W m2
(Bond et al., 2011). The UNEP/WMO assessment
reports a central net value of BC and OC forcing
(including the snow albedo efficacy factor) of 0.41 W
m2 (UNEP and WMO, 2011a).
Several additional factors must be taken into
consideration in interpreting these estimates. First,
it is important to note that like BC, OC exhibits
high spatial variability in direct forcing effects (see
Figure 2-18). The regions of highest direct forcing
by OC may not coincide with regions of highest
direct forcing by BC (see Figures 2-13 and 2-15 for
comparison). In addition, most studies evaluating
the net effects of BC and OC do not consider
indirect effects, and inclusion of these effects will
change the net forcing estimates. One study, Chen
et al. (2010), found that for one scenario reducing BC
and OC in a 2 to 3 ratio, the aerosol indirect effects
were larger than (and opposite in sign to) the direct
effects. In addition, studies looking at forcing effects
due to OC generally consider primary OC emissions
only. Secondary organic aerosols (SOA), however,
can also make a substantial contribution to the
organic aerosols. SOA arises from the oxidation of
gaseous VOCs. More recently, Robinson et al. (2007)
proposed a more dynamic evolution of aerosol OC
in the atmosphere. Based on measurements and
models, they suggested that low volatility organic
compounds, which are emitted as PM, evaporate,
oxidize, and condense over time. The semi-volatile
nature of the primary emission of OC may have
additional implications for our understanding of
OC and OC to BC ratios on climate, although this
remains poorly understood (Jimenez et al., 2009).
The inclusion of other species, mainly nitrate and
sulfate aerosols, also tends to reduce the estimate
of net forcing. In particular, the presence or absence
-90
-180
-90
0
90
Figure 2-18. Direct Forcing by OC from All Sources, simulated with the Community Atmosphere Model.
(Bond etal., 2011)
44
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Black Carbon and Its Effects on Climate
of sulfates and nitrates, which
together comprise a large
fraction of aerosol mass, in
calculations of indirect effects
can dominate radiative forcing
calculations. Inclusion of both
direct and indirect effects of
aerosol species in the review by
Ramanathan and Carmichael
(2008) led to an estimate of the
total aerosol effect including
direct and indirect effects of
-1.4 W rrv2, in contrast to a
calculated BC direct forcing of
+0.9 W m2. However, because
much of the nitrate and sulfate
precursor emissions come from
sectors that are not rich in BC,
the net global effect of aerosols
can be less important than the
estimates of the net effects of
aerosols from a specific sector
or measure (discussed further in
section 2.6.1.6). These aerosols
also play a role in the mixing
state and therefore the direct
radiative forcing effect of BC,
as discussed in sections 2.5
and 2.6.1.2. Therefore, ambient
concentrations of these other
aerosols can be important in
determining the influence of BC
reductions. Using surface and
aircraft measurements, Ramana
et al. (2010) found that the ratio
of BC to sulfate was important
in determining the net warming
or cooling impact of pollution
plumes in China.
2.6.1.6 Sector-Based
Contributions to Radiative
Forcing
(3) On-road(199)
Household biofuel (132)
Animal Husbandry (98)
Household fossil fuel (84)
Waste/landfill (84)
Power (79)
Agriculture (29)
Off-road land (20)
Aviation (-6)
Agr. waste burning (-14)
Shipping (-43)
Biomass burning (-106)
Industry (-158)
(b)
Power (554)
On-road(417)
Industry (283)
Household fossil fuel (254)
Household biofuel (159)
Animal Husbandry (131)
Agriculture (98)
Waste/landfill (88)
Off-road land (39)
Aviation (27)
Biomass burning (22)
Agr. waste burning (-14)
Shipping (-22)
ii
Ozone
Sulfate
Nitrate
Black carbon
Organic carbon
AIE
Methane
Nitrous Oxide
Carbon Dioxide
-600 -400 -200 0 200 400
Radiative forcing (mWrrr2)
600
I
Ozone
Sulfate
Nitrate
Black carbon
Organic carbon
AIE
Methane
Nitrous oxide
Carbon dioxide
As described in Chapter 4, BC
emissions can be attributed to
a wide range of sectors (e.g.,
transportation, residential,
industrial, and biomass burning).
Some studies have attempted
to quantify the radiative forcing
effect of emissions mixtures containing BC and
other co-pollutants by estimating the radiative
forcing of defined emissions sectors. Comparisons
among studies, however, is hindered in part by
variation in several parameters, including estimates
of the sector-level contributions, the relative
-600 -400 -200 0 200 400 600 800 1000
Radiative forcing (mWm-2)
Figure 2-19. Global Radiative Forcing Due to Perpetual Constant Year 2000
Emissions, Grouped By Sector, at (a) 2020 and (b) 2100 and Showing the
Contribution from Each Species. The sum is shown on the title of each bar,
with a positive radiative forcing means that removal of this emissions source
will result in cooling. AIE is the aerosol indirect effect. (Linger et al., 2010)
fraction of warming and cooling aerosols, and the
microphysical properties of these aerosols.
Unger et al. (2010) examined the impacts of
sector-specific emissions on the short- and
long-term radiative forcing from a range of
pollutants. Figure 2-19 shows that the mixture of
Report to Congress on Black Carbon
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Chapter 2
emissions from some of the largest BC emissions
sources contributes considerably to total radiative
forcing. On-road transportation emissions are
the largest contributor to radiative forcing in the
short term (by 2020), due to a combination of GHG
and BC emissions. On-road transportation is also
seen to be the second largest contributor in the
long term (by 2100), but this is largely the result
of the significant GHG emissions from this sector.
Residential biofuel combustion is the second largest
contributor in the short term due to the contribution
from BC and CH4. Since these sources have lower net
GHG emissions, they contribute less to total global
radiative forcing in the long term. However, these
calculations have substantial uncertainties owing
to the details of aerosol physics and chemistry, the
interactions of aerosols and clouds, and the regional
nature of the radiative forcing, as discussed earlier
in this chapter. Bauer and Menon (2012) focused
on regional differences in the impact of emissions
from different source categories, and concluded that
the largest opportunities to reduce positive forcing
due to all aerosols included transportation in all
regions, agricultural burning in Europe and Asia, and
residential cooking and heating ("domestic sector")
in Asia.
There is significant disagreement regarding the net
impact of aerosol emissions from open biomass
burning on radiative forcing. As noted in the
previous section, the IPCC estimated the net direct
radiative forcing impact from open biomass burning
aerosols to be small, but positive at +0.03 (±0.12)
W m2 (Forster et al., 2007). However, because of
uncertainties regarding the extent and composition
of emissions from this source category, and
the indirect radiative forcing effects of biomass
burning aerosols, it is not clear if this sector has an
overall global warming or cooling effect. Kopp and
Mauzerall (2010) developed probability distributions
from multiple studies to examine the likelihood of
warming from individual sectors. Based on existing
evidence, they concluded that open burning in
forests and savannas is unlikely to contribute to
warming, while the effect of open burning of crop
residues remains uncertain. The results of current
analyses are sufficiently different that there is no
consensus on the likelihood of warming. Stohl et
al. (2007) concluded that biomass burning has a
"significant impact on air quality over vast regions
and on radiative properties of the atmosphere" and
in particular "has been underestimated as a source
of aerosol and air pollution for the Arctic, relative to
emissions from fossil fuel combustion." As discussed
further in section 5.6, surface snow records indicate
that biomass burning is currently a major source of
BC in Greenland and the North Pole (Hegg et al.,
2010). Bauer and Menon (2012) point to agricultural
burning in Europe and Asia as contributing to
positive net forcing in those regions. Additional
work is needed to improve scientific understanding
of the radiative forcing impacts of open biomass
burning.
Several modeling experiments, such as Jacobson
(2002, 2005, 2010), Hansen et al. (2005), and Schulz
et al. (2006), and observationally constrained
studies such as Ramanathan and Carmichael (2008),
have found that carbonaceous aerosols from
biofuel combustion and fossil fuel combustion
both contribute to warming. Among fossil fuels,
diesel combustion for transportation is the largest
contributor to global BC emissions and several
studies suggest these emissions may contribute
to warming (Jacobsen 2002, 2005; Hansen et
al., 2005). Kopp and Mauzerall (2010) concluded
that carbonaceous PM emissions from gasoline
combustion are unlikely to contribute to warming,14
while diesel combustion and residential coal
combustion are very likely to contribute to warming.
Kopp and Mauzerall (2010) also found mixed results
with respect to the contribution of residential
biofuel combustion for the models included in their
assessment. More recent work by Rehman et al.
(2011) has reported higher BC concentrations due
to burning biomass for cooking than previously
reported, both indoors and outdoors, in the study
region in northern India. Moreover, Rehman et al.
found that the albedo of the particles in the study
villages indicated high absorption. Wavelength
analysis suggested that though OC concentrations
were a factor of five higher than BC concentrations,
the OC included significant absorbing BrC.
Aviation is also a source of BC emissions. While the
amount of BC emitted by aircraft at cruise altitudes
is subject to large uncertainties (see Appendix 2-10),
BC particles at these altitudes absorb not only the
downward radiation but also the reflected upward
radiation. In addition, intercontinental flight tracks
are concentrated in the arctic stratosphere and
particles emitted in this region may be deposited
in the Arctic ice and snow. Research is needed to
quantify radiative forcing due to BC of aviation
origin.
A few studies highlight the substantial uncertainties
regarding the contribution of biofuel combustion
14 For gasoline vehicles, it should be noted that the
introduction of new engine technologies (e.g., some types of
gasoline direct injection) in recent model years has increased
BC/PM ratios in some new gasoline-powered motor vehicles
(Smallwood et al., 2001), which may change the warming
profile of emissions from these vehicles. See Chapter 8 for
more discussion of this issue.
46
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Black Carbon and Its Effects on Climate
and fossil fuel combustion to warming, given
our limited understanding of how carbonaceous
aerosols affect cloud processes. In the modeling
experiments by Chen et al. (2010), reductions in
fossil fuel carbonaceous aerosols (BC and OC)
lead to decreases in CCN, leading to a decrease
in cloud albedo, causing an increase in radiative
forcing. The impact of these cloud changes equal
or exceed the direct radiative forcing impacts. This
result contrasts with that of Jacobson (2010) and
Bauer et al. (2010) in which estimated warming
from indirect effects did not exceed the direct and
other radiative forcing from fossil fuel emissions.
The distinction between biofuel combustion and
fossil fuel combustion, in terms of the effect on
radiative forcing, is particularly noticeable in the
residential sector. Aunan et al. (2009) estimate that
the global annual mean radiative forcing from BC
from residential fuel consumption in Asia is positive
for both biofuel and fossil fuel consumption, but that
the net effects on radiative forcing from residential
biofuel consumption in Asia (accounting for BC and
the range of other co-emitted pollutants) is negative.
Fossil fuels burned for electricity generation
contribute only a small fraction of carbonaceous
aerosol emissions, though this sector is a large
source of long-lived, warming GHGs and short-lived
cooling sulfate aerosols (Shindell and Faluvegi, 2009).
Thus, though their study found that the sector is the
largest single contributor to warming on the 100-
year time scale, this is attributable to GHG emissions
rather than emissions of BC.
2.6.2 Impact of BC Radiative Forcing on
Temperature and Melting of Ice and Snow
As mentioned in section 2.6.1, global average
radiative forcing is linearly related to the global
mean temperature at the surface (Forster et al.,
2007). Radiative forcing from agents such as BC
has similar effects on global mean temperature as
radiative forcing from CO2 and other GHGs (Hegerl
et al., 2007), though the efficacy of the forcing may
differ slightly (especially for the snow/ice albedo
effect). Temperature itself has already been linked
to a range of climate impacts as identified in, for
example, the 2009 USGCRP report, "Global Climate
Change Impacts in the United States." This and
other recent climate change assessments describe
the risks and impacts associated with climate
change, including degradation of air quality,
temperature increases, changes in extreme weather
events, effects on food production and forestry,
effects on water resources, sea level rise, disruption
to energy consumption and production, and
potential harm to ecosystems and wildlife. Though
few studies explicitly link BC to all of these outcomes,
to the extent that BC increases temperature it will
contribute to these impacts, especially impairment
of air quality and sea level rise (via melting of ice,
snow, and glaciers).
Work by Jacobson often uses temperature change
as an endpoint, rather than radiative forcing. As
discussed earlier in section 2.6, there are advantages
and disadvantages to using temperature change
as an endpoint. Jacobson also includes a more
complete suite of BC and co-pollutant effects than
most other models. Jacobson (2010) found that
the net effect of existing fossil fuel BC plus OC
emissions was to warm the climate by 0.3 to 0.5°C
at equilibrium compared to a case without those
emissions.
There have been some efforts to translate regional
direct radiative forcing estimates into regional
changes in temperature. For example, Chung and
Seinfeld (2005) used the GISS GCM model with BC
emissions from Bond et al. (2004) to predict annually
averaged changes in regional temperatures due to
BC direct radiative forcing, based on simulations
of 100 years of future emissions and temperature
responses. They predict that externally mixed BC
leads to an increase in average annual surface
temperatures of 0.29°C in the Northern Hemisphere
and 0.11°C in the Southern Hemisphere, when
model results for the latter 75 years of simulations
are averaged (the first 25 years of simulations were
omitted from the results, to allow for the model to
settle). Internally mixed BC is predicted to result in
an increase in average annual surface temperatures
of 0.54°C in the Northern Hemisphere and 0.20°C
in the Southern Hemisphere over the same
period. Despite the uncertainties regarding mixing
state, the authors show that surface temperature
response to regional direct radiative forcing is more
concentrated in the Northern Hemisphere, especially
in locations of high latitude.
Few studies have evaluated the North
America-specific temperature impacts associated
with BC emissions. However, Qian et al. (2009) found
that BC emissions lead to warming of a tenth to a
full degree Celsius over snow in the western United
States. Simulations show that BC absorption of solar
radiation in the atmosphere leads to as much as
0.6°C of warming in the lower and mid troposphere
over most of North America, including the Arctic
region (Ramanathan and Carmichael, 2008).
Additional estimates of regional temperature effects
associated with BC emissions in the Arctic and the
Himalayas are discussed in sections 2.6.4 and 2.6.5,
respectively.
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Chapter 2
The snow/ice albedo effects of BC have been linked
to accelerated melting of snow and ice. While many
glaciers around the world and Arctic sea ice have
receded in recent decades, attribution of this melt
to BC is challenging due to other global and local
contributions to warming and precipitation changes.
Regardless of the deposited BC, other factors -
including the solar zenith angle, cloud cover, snow
grain size, and depth of the snow - also influence
the snow and ice albedo (Wiscombe and Warren,
1980). The most common method of determining the
contribution of BC to accelerated snow and ice melt
has been to compare model runs with and without
BC influences, and evaluate with observations. Direct
measurements are generated by melting and then
filtering samples of snow and ice. The filters provide
an estimate of BC concentration by comparing
their observed optical transmissions to optical
transmissions of known amounts of BC (Noone and
Clarke, 1998; Warren and Clarke, 1990). The mass
is then used to estimate or compare to measured
snow albedo, calculating the influence of BC. Another
approach has been to apply a known amount of soot
to an area, and then compare the measured albedo
and melting rate to a nearby clean plot of snow.
In the western United States, deposition of BC on
mountain glaciers and snow packs produces a
positive snow and ice albedo effect, contributing to
the melting of snowpack earlier in the spring and
reducing the amount of snowmelt that normally
would occur later in the spring and summer (Hadley
et al., 2010; Koch and Del Genio, 2010). This has
implications for freshwater resources in regions of
the United States that are dependent on snow-fed
or glacier-fed water systems. In the Sierra Nevada
mountain range, Hadley et al. (2010) found BC at
different depths in the snowpack, deposited over
the winter months by snowfall. In the spring, the
continuous uncovering of the BC contributed to the
early melt. A model capturing the effects of soot on
snow in the western United States shows significant
decreases in snowpack between December and
May (Figure 2-20, Qian et al., 2009). Snow water
equivalent (the amount of water that would be
produced by melting all the snow) is reduced
by 2-50 millimeters (mm) in mountainous areas,
particularly over the Central Rockies, Sierra Nevadas,
and western Canada. In addition, dust deposition
on snow, at high concentrations, can have similar
effects to BC. A study done by Painter et al. (2007)
in the San Juan Mountains in Colorado observed
a decrease in snow cover duration of 18-35 days
as a result of dust transported from non-local
desert sources. As the authors note, in the future,
exacerbated dryness in desert and arid regions—
SWE change (mm)
Mar
50
20
10
5
2
-2
-3
-10
-20
-50
32N-
129W 126W 123W 120W 117W 114W 111W 108W 105W
Figure 2-20. Spatial Distribution of Change in Mean Snow Water Equivalent (SWE, mm) for March. (Qian et al., 2009)
48
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Black Carbon and Its Effects on Climate
especially in the Southwest United States, the
Middle East, and the Sahel—could cause increased
deposition of dust on mountain snow cover in
areas such the Rocky Mountains, the HKHT, and the
Alps, respectively, which could lead to considerable
reductions in snow cover duration over large areas.
Changes in snow and ice melt due to BC can affect
various types of surfaces and geographic locations
throughout the world, including Arctic ice caps
and sea ice, glaciers, and mountain snowpack (see
section 2.6.4 for more detailed treatment of Arctic
impacts). For example, Ming et al. (2009) suggest
that reduced albedos in some glaciers in west China
from BC deposition might accelerate melting of these
glaciers. Figure 2-21 shows a Chinese glacier and the
concentration of BC that results from melting the
upper layers of the snowpack until it is buried by
fresh snowfall.
It is important to note that the impacts of BC on
snow and ice albedo are not constrained to regions
of high elevation or high latitude. BC deposition can
also contribute to accelerated melting of seasonal,
non-mountain snow, especially in mid-latitude
regions due to the additional exposure to sunlight
these regions experience (compared to polar
regions). In addition, these regions are generally
closer to BC emissions sources than polar regions
are, so concentrations of BC on seasonal snow
accumulations can be considerable (Huang et al.,
2011).
2.6.3 Other Impacts of BC
In addition to warming and cooling effects due to
absorption and reflection of light both directly and
through cloud interactions, BC and other aerosols
contribute to climate change through surface
dimming and changes in precipitation patterns.
The following sections provide information on BC's
contributions to these two types of impacts. It is
important to note that unlike BC and other aerosols,
GHGs are not associated with surface dimming, nor
are they linked directly to changes in precipitation.
Changes in precipitation from GHGs are mediated
through changes in temperature.
2.6.3.1 Surface Dimming Effects
The absorption of incoming solar radiation by BC
reduces the amount of solar radiation reaching
the Earth's surface, an effect referred to as surface
dimming in many studies (e.g., Forster et al.,
2007). This results in cooling at the surface (even
though net forcing measured at the TOA may be
positive). A number of studies report evidence
of global dimming between the 1960s and the
1980s, followed by an increase in the amount of
sunlight reaching the surface during the 1990s to
(a)
600
400
O 200
CO
ZD snow pit of 2006
Summer
Dust layer
I L
10 20
Depth (cm)
30
40
(b)
o
"o
CO
Figure 2-21. BC Concentrations in
the ZD Glacier, (a) Measured BC
concentration (b) the snow pit of the ZD
Glacier. Dust layer at 30 cm indicated the
spring (melting season) of 2006. (From
Ming et al., 2009, Figure 4)
Snow pit
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Chapter 2
the present (sometimes referred to as brightening]
(e.g., see review in Wild, 2009). Numerous studies
suggest that the observed dimming and brightening
trends are caused by changes in aerosol emissions
over time and the interaction of aerosol direct and
indirect radiative forcing (Stanhill and Cohen, 2001;
Wild et al., 2005; Streets et al., 2006; Ruckstuhl et al.,
2008). See Table 2-5 for a summary of how aerosol
interactions affect surface dimming.
Estimating the magnitude of the surface dimming
effect is complicated, largely because the vertical
distribution of BC in the atmospheric column
affects the impacts at the surface. Isolating the
effect of BC is also difficult due to the interactions
of non-BC aerosols. For example, many non-BC
aerosols (primarily sulfates) scatter incoming solar
radiation, reducing the energy reaching the surface
(Dwyer et al., 2010; Ramanathan and Carmichael,
2008). In addition, the indirect effect of aerosols
on cloud albedo and cloud lifetime may decrease
solar radiation at the surface (Ramanathan and
Carmichael, 2008). Furthermore, surface cooling
combined with atmospheric heating from BC may
increase the stability of the boundary layer (e.g., the
bottom layer of the troposphere that is in contact
with the surface of the earth) and reduce vertical
mixing. This increase in atmospheric stability reduces
natural removal processes for air pollutants, resulting
in worse air pollution episodes (Ramanathan and
Carmichael, 2008).
In some regions, BC, BrC, sulfates, organics, dust
and other components combine to form pollution
clouds known as ABCs, which have been linked
to global dimming (Ramanathan and Carmichael,
2008; Ramanathan et al., 2007). Ramanathan and
Carmichael (2008) estimate the total global dimming
effect from ABCs to be -4.4 W rrr2, with about
-3.4 W rrr2 from the direct effect of aerosols (roughly
half of which is attributed to BC) and the remaining
-1 W rrr2 from the indirect effect. In the Ramanathan
and Carmichael study (2008), the -1.7 W m 2 of
surface dimming from BC was found to be offset
by +2.6 W m"2 of heating in the atmosphere. This
resulted in a net TOA forcing estimate from this
study of +0.9 W m 2, as cited in section 2.6.1.2.
The dimming effects due to BC and the other
aerosols are not spatially uniform (see Figure 2-22). A
number of studies have found that dimming effects
are particularly acute in certain regions associated
with high aerosol pollution levels and the presence
of ABCs. These include certain major urban areas
(Ramanathan and Feng, 2009; Trenberth et al., 2007),
primarily in South Asia (Ramanathan and Feng,
2009; Ramanathan et al., 2005; 2007). The ABCs
which cover large areas in the North Indian Ocean
and South Asia can reduce energy at the surface by
5-10% (Ramanathan et al., 2007; Ramanathan and
Carmichael, 2008). Some studies have estimated
that the dimming associated with ABCs can mask
approximately half of the warming that would occur
at the surface in the absence of ABCs due to GHGs
alone, especially over Asia (Ramanathan et al., 2007;
UNEP, 2008a). Another study estimates that surface
dimming causes a reduction of approximately 6% in
solar radiation at the surface over China and India
when compared to pre-industrial values (UNEP,
2008a). The U.S. Global Change Research Program
(CCSP, 2009) estimated surface forcing values as
low as -10 W m 2 over China, India, and sub-Saharan
Africa due to elevated optical depth from aerosol
emissions in that region.
2.6.3.2 Precipitation Effects
Aerosols affect the processes of cloud and rain
droplet formation. Some studies have linked
aerosols to reductions in rainfall, but these
interactions are not well understood. A summary
of the aerosol interactions with clouds that cause
Table 2-5. Overview of the Different Aerosol Indirect Effects and Their Implications for Global Dimming
and Precipitation. This table applies to all aerosols, not just BC. Scientific uncertainty is "very low" for all
effects except the cloud albedo effect (for which uncertainty is "low"). For descriptions of the effects, see
section 2.6.1.3. (Adapted from Denman et al., 2007, Table 7.10b.)
Fff Sign of Change in Potential Sign of Change in Potential
Surface Dimming Magnitude Precipitation Magnitude
Cloud albedo effect
Cloud lifetime effect
Semi-direct effect
Glaciation indirect effect
Thermodynamic effect
Positive
Positive
Positive
Negative
Positive or negative
Medium
Medium
Large
Medium
Medium
NA
Negative
Negative
Positive
Positive or negative
NA
Small
Large
Medium
Medium
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Black Carbon and Its Effects on Climate
-20 -12
-3 -1
Figure 2-22. Surface Dimming by Anthropogenic Aerosols (W m2). (Adapted from Chung et al., 2005,
and Ramanathan and Carmichael, 2008)
changes in precipitation is provided in Table 2-5.
According to the IPCC (2007), the precipitation
effects on a global scale attributed primarily to BC
are from the semi-direct effect (described in section
2.6.1.3) and the increased atmospheric stability
resulting from that effect (Ramanathan et al., 2005;
Chung and Zhang, 2004; Menon et al., 2002). The
increased stability inhibits convection, affecting both
rainfall and atmospheric circulation. As discussed
in section 2.6.1.3, increased availability of CCN
increases cloud lifetime, thereby inhibiting rainfall
for a time period, which may be more important for
shifting the location of rainfall than changing net
global precipitation. There may also be increases
in precipitation: BC in particular can stimulate
precipitation from ice clouds. However, because
of the dependence of precipitation on complex
and localized conditions, scientific understanding
of these effects is low and models often disagree
on the magnitude or sometimes even the sign of
changes in precipitation due to factors such as
warming or aerosol emissions.
Surface dimming due to all types of aerosols
may reduce precipitation by reducing the energy
available for evaporation from the Earth's surface
(Liepert et al., 2004; Ramanathan et al., 2001). Bauer
and Menon (2012) estimate that the global mean
change in precipitation due to all aerosols is -0.3%.
Because rain is a major removal mechanism for BC
from the atmosphere, large decreases in rainfall
could result in higher atmospheric concentrations of
BC and other aerosols (Ramanathan and Carmichael,
2008; Ramanathan et al., 2005).
Ramanathan and Feng (2009) suggest that, on a
global average basis, reduced precipitation caused
by the surface dimming effects of aerosols is likely
to be countered with increased precipitation from
GHG-induced warming. The effect of aerosols on
precipitation, however, varies by area, surface cover,
and location. For example, in the tropics, the net
effect of aerosols and GHG-induced warming may
be reduced rainfall (Ramanathan and Feng, 2009).
These shifts in rainfall patterns may have important
implications for water availability.
In the United States, Qian et al. (2009) found only
small changes in the amount of precipitation in
the western U.S. as a result of BC effects. While
there is no evidence in North America that links
BC or any other specific constituent of PM to
changes in precipitation, there are studies that
show correlations between total PM emissions
and regional precipitation patterns. For example,
Bell et al. (2008) find weekly patterns of emissions
that correlate with weekly patterns in rainfall in the
southeastern United States (Bell et al., 2008). Similar
results have also been found for the East Coast of
the United States (Cerveny and Balling, 1998).
There is stronger evidence linking aerosols to
reduced precipitation in the tropics. Studies have
indicated that surface dimming in this region
reduces evaporation (Feingold et al., 2005; Yu
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Chapter 2
et al., 2002a; Hansen et al., 2007b). Other studies
have found that the effect in the tropics may be
unevenly distributed with increased precipitation
just north of the equator (between 0° and 20°
N) and decreased precipitation just south of the
equator (between 0° and 20° S). This would shift the
Intertropical Convergence Zone northward (Chung
and Seinfeld, 2005; Roberts and Jones, 2004; Wang,
2004). This northward shift may be caused by the
enhanced temperature difference between the
Northern and Southern Hemispheres (also a result
of global warming from GHGs), which induces a
change in circulation and convection in the tropics.
Aerosols have also been linked to impacts on
regional precipitation in the Amazon basin (Martins
et al., 2009; Bevan et al., 2009). This is a region of
high biomass burning emissions in the dry season.
Further, seasonal biomass emissions have been
linked to larger changes in atmospheric circulation
patterns by affecting the global distribution of high-
level clouds and convection precipitation (Jeong
and Wang, 2010). Jeong and Wang (2010) also found
that the climate response extends outside of the
biomass burning season. The effects of BC aerosols
on precipitation may also extend beyond areas of
high concentrations. Wang (2007) found the largest
change in precipitation occurs in the tropical Pacific
region which is far from the regions of largest BC
forcing. The effect may be very similar to the pattern
of precipitation anomalies associated with the El
Nino/Southern Oscillation.
There is also evidence that BC and ABCs slow down
the monsoon circulation over South Asia. Specifically,
the surface dimming caused by BC aerosols
(Meehl et al., 2008) and ABCs (Lau and Kim, 2006;
Ramanathan et al., 2005) alters both the north-
south gradients in sea surface temperatures and the
land-ocean contrast in surface temperatures. These
studies estimate an increase in pre-monsoon rainfall
during spring followed by a decrease in summer
monsoon rainfall, in agreement with observed
trends.
Model studies of China have found that BC
contributes to increased rainfall in the south and
reduced rainfall in the north (Wu et al., 2008; Menon
et al., 2002). Wu et al. (2008) simulated the regional
climate impacts of BC's direct radiative forcing
effect in Asia and found about a 0.6% increase
in atmospheric water vapor over southern China,
resulting in a precipitation increase of 0.4-0.6 mm/
day. In northern China, this study found about a 0.3%
decrease in water vapor and a resultant decrease
in precipitation. Meehl et al. (2008) found small
precipitation increases over the Tibetan Plateau due
to BC, but concluded that precipitation over China
generally decreases due to BC effects.
2.6.4 BC Impacts in the Arctic
BC emissions that are transported to the Arctic
are strongly linked to local warming (Reddy and
Boucher, 2007), even if the globally averaged net
climate impact of the total particulate emissions
from individual sources is uncertain. For example,
Quinn et al. (2008) calculated that the contribution
of short-lived climate forcers (i.e., CH4, tropospheric
ozone, and tropospheric aerosols, including BC)
to Arctic warming is about 80% that of CO2. BC
can have significant snow albedo effects and the
magnitude of the cooling effect over snow from
co-emitted aerosols is reduced in the Arctic.
Because temperature in the Arctic has warmed
at twice the global rate over the past 100 years
(IPCC, 2007) and because of the dramatic retreat
of summer sea ice extent during the satellite
observation period (see Figure 2-23), there is interest
in mitigation strategies that may slow the near-term
rate of climate change in this region.
The estimated radiative forcing from BC is larger
over the Arctic than it is on average globally. Due
to the lack of sunlight in winter months, the long
days in summer, and the increased efficiency of
transport of BC emissions from lower latitudes in
spring, there is also much larger seasonal variability
in the estimates of radiative forcing from BC and
other aerosols than there is from GHGs (Quinn et
al., 2008). Looking at forcing from fossil fuel and
biofuel BC emissions, Quinn et al. (2008) calculated
a radiative forcing in the Arctic of +1.2 W rrr2 in the
spring, +0.66 W m 2 in the summer, +0.16 W m2
in the fall, and only 0.09 W rrr2 in the winter. Snow
albedo forcing in the Arctic was calculated to add
an additional +0.53 W rrv2 in the spring, +0.21 W
m^2 in the summer and negligible forcing in autumn
and winter. This effect is amplified (e.g., increase
in efficacy) by the hastening of the spring thaw
that reveals darker ground and water (ocean/lake)
surfaces.
Due to the frequency of strong temperature
inversions that inhibit atmospheric mixing, and
the prevalence of dry conditions that impede wet
deposition, the lifetime of aerosol particles in the
Arctic is longer than other regions - sometimes
weeks, rather than days (see Garrett et al., 2004;
Curry, 1995). This leads to a phenomenon known as
Arctic haze which is the result of an accumulation
of BC, OC, and sulfate particles in the atmosphere
above the Arctic (Quinn et al., 2007). Strong
surface-based temperature inversions and the
dryness of the Arctic troposphere inhibit removal
of particles via deposition. Over a highly reflective
surface like the Arctic, BC particles absorb solar
radiation and warm the atmosphere above
52
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Black Carbon and Its Effects on Climate
1981 -2000 average
2007
Second-year ice
(1-2 years old)
Older ice • Open
(>2 years old) water
Land
2007
Figure 2-23. Evidence of Arctic Ice Melt, (a) Extent of summer Arctic sea ice for 2007-2009 compared to 1981-
2000 average. (Source: National Snow and Ice Data Center, http://nsidc.org/news/press/20091005_minimumpr.
html) (b) Duration of summer surface melt on Greenland in 2007 relative to 1973-2000 average (Source:
Arctic Monitoring and Assessment Programme, 2009). Arctic summer sea ice has decreased by 40% since
1979, accompanied by in increasing discharge from the Greenland ice sheet. Natural variability may explain
some of these changes, but the overall trend toward warming and melting has been attributed primarily to
human-induced climate change (Min et al., 2008; Holland et al., 2008). Summer sea ice melt creates a feedback
loop that amplifies warming as reflective white ice/snow surfaces are replaced by darker ocean waters,
increasing sunlight absorption. Recent work suggests a link between Arctic sea-ice melt and increased glacier
runoff in Greenland (Rennermalm et al., 2009).
and within the haze layer, while simultaneously
contributing to surface dimming. Rather than a
cooling effect from surface dimming, however,
the atmospheric heating increases the downward
longwave radiation and causes warming at the
surface (Shaw and Stamnes, 1980; Quinn et al., 2008;
Mauritsen et al., 2011). Any warming particle above
a highly reflective surface can lead to heating of
the entire surface-atmosphere aerosol column. In
addition, the stable atmosphere above the Arctic
prevents rapid heat exchange with the upper
troposphere, increasing surface warming in the
Arctic (Hansen and Nazarenko, 2004; Quinn et al.,
2008).
Radiative forcing from both atmospheric
concentration and deposition on the snow and ice
has contributed to the surface temperature warming
in the Arctic (Quinn et al., 2008). Simulations by
Planner et al. (2007) suggest that the deposition
of BC from sources in North America and Europe
on Arctic sea ice may have resulted in a surface
warming trend of as much as 0.5 to 1°C. Similarly,
Shindell and Faluvegi (2009) found 0.5 to 1.4°C of
warming from BC in the Arctic since 1890. For the
BC snow albedo effect, Quinn et al. (2008) estimated
a warming of 0.24 to 0.76°C, varying by season.
Warming due to BC heating in the atmosphere is
estimated to be a further 0.24°C in spring, 0.15°C
in summer, and nearly zero in autumn and winter.
In Table 2-6, we show estimates of temperature
increases in the Arctic from various BC emissions
sources (Shindell and Faluvegi, 2009; Flanner et al.,
2007; Jacobson, 2010). Part of these increases in
temperature may also have been "unmasked" in
recent years from reductions in sulfate aerosols and
its gaseous precursor, SO2 (Shindell and Faluvegi,
2009). While sulfate aerosols have a negative
Report to Congress on Black Carbon
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Chapter 2
Table 2-6. Arctic Temperature Impacts from Emissions of BC from Different Sectors.
PhySl^U SSSS5S2S ~ A"°'°" M°<*' "•*»-
Arctic
Arctic
Arctic
Arctic
Arctic
0.5 -1. 4° C
0.5 -1. 6° C
1.2°C
1.7°C
Indirect > 0
100%ofFF,
BF, BB
100%ofFF,
BF, BB
100%ofFF
100%ofFF, BF
100% of All
sources
BC.OC
BC.OC
BC.OC, minor
inorganics
Same as
above plus
C02, CH4
All aerosols
GISS-ER
NCAR-CAM3
andSNICAR
GATOR-
GCMOM
Same as
above
Idealized
calculations,
observations
Shindelland
Faluvegi (2009)
Planner etal.
(2007)
Jacobson(2010)
Jacobson(2010)
Mauritsen etal.
(2010)
Aerosol indirect
included "crudely."
Range results
from using fire
frequencies in high
year globally (1998)
and low year (2001).
Indirect only, no
direct or snow
albedo effect.
FF = fossil fuel; BF = biofuel; BB = biomass burning.
radiative forcing, the reductions in sulfate aerosols
have been strongly justified by improvements in
air quality, acid rain, visibility, public health, and
lessening of direct effects of sulfates on ecosystems.
It has also been suggested that the potential cooling
effects of BC—such as indirect radiative forcing and
the ratio of BC to cooling components (e.g., OC)—
may not be as important in the Arctic since the snow
and ice albedo darkening is so dominant (Mauritsen
et al., 2010).
While there are strong qualitative indications of
Arctic snow and ice melt from BC, quantitative
studies have only recently entered the peer-reviewed
literature. Some studies have linked the local
warming measured on the Greenland Ice Sheet to
observations of a gradual loss of ice, and modeled
the overall impact on the mass balance of the ice
sheet. Box et al. (2004), for example, estimated the
modeled ice sheet mass balance at -76 km3 per
year, leading to a 0.24 mm sea level rise per year
(contributing 15% of global sea level rise) during
1991-2000. Hanna et al. (2005) considered a longer
time period, and estimated that the overall mass
balance declined at a rate of -22 km3 per year in
1961-1990 and -36 km3 per year for 1998-2003, with
melting during the past six years contributing 0.15
mm per year to global sea level rise. Finally, Thomas
et al. (2006) reported accelerating mass loss between
an earlier period (between 4 and 50 Gigatons (Gt)
per year, depending on the model, from 1993-1999)
and a more recent period (between 57 and 105 Gt
per year, from 1999-2004). In a modeling study by
Planner et al. (2007), land snowmelt rates north of
50°N latitude (about 70 miles north of the U.S./
Canada border in Minnesota) increased by 28%
in 1998 and 19% in 2001 in the month preceding
maximum melt when compared to control runs
that did not include BC from large boreal fires that
occurred in 1998 and 2001. Strack et al. (2007) found
soot deposition in the Alaskan Arctic tundra created
snow-free conditions five days earlier than model
runs without BC deposition. Ongoing studies will
help evaluate and constrain modeling simulations.
Importantly, American, Norwegian, Russian, and
Canadian research groups collaborated under the
International Polar Year (2007-2008) program to
survey BC concentrations in snow and ice north of
65°N latitude in both the Eastern and Western Arctic
(Doherty et al., 2010).
The location of the emissions also matters for the
magnitude of the effects in the Arctic, which has
important implications for mitigation decisions.
A recent study by the Arctic Monitoring and
Assessment Program (Quinn et al., 2011) analyzed
the radiative forcing impacts of BC emissions from
different regions on the Arctic (direct and snow/ice
albedo only). This study found that compared to the
average emissions of BC from regions between 40°N
and 50°N latitude, emissions of BC from between
50°N and 60°N latitude had about three times as
much forcing impact in the Arctic on a per-ton basis.
In addition, emissions from north of 60°N had seven
times as much impact per ton. However, because
total emissions are much larger in the southern
regions, almost half of the total impact on Arctic
forcing due to BC in this study was derived from
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Black Carbon and Its Effects on Climate
emissions from regions below 50°N, and most of the
remainder was derived from emissions from regions
between 50°N and 60°N.
BC emissions from near-Arctic countries have
decreased since their peak in the early 20th century.
This is supported by a downward trend in the
observed concentrations of ambient and snow/
ice BC in the Arctic (see section 5.6). However, BC
deposited on areas covered permanently with ice
and snow, such as the Greenland Ice Sheet, can
remain in the ice sheet for hundreds of years, as
seen in historical ice core records (Quinn et al., 2008).
Although the sunlight that reaches the snow surface
typically only penetrates between 10 and 20 cm
deep, with the topmost 5 cm of snow receiving the
most sunlight (Galbavy et al., 2007), as the Arctic
warms and snow and ice melt, deeper, hidden BC
that was deposited over decades may become
exposed, enhancing the melting of snow and ice.
The effect of BC on the snow and ice albedo in the
Arctic thus can involve historical—in addition to
present-day—BC deposition in the Arctic region.
An important uncontrolled source of BC in some
near-Arctic countries is open biomass burning.
Several recent studies have looked at the effect of
these emissions on the Arctic. For example, Stohl
et al. (2006) found that North American boreal
forest fires lead to elevated concentrations of
light absorbing aerosols including BC throughout
the entire Arctic, with substantial implications for
Arctic warming and enhanced snow albedo effects.
Analyses conducted by Hegg et al. (2010) suggest
that the dominant source of light-absorbing aerosols
(including BC) in the Arctic region is biomass
burning. Other studies have linked open biomass
burning to reduced surface albedo and accelerated
melting (Hegg et al., 2009; Generoso et al., 2007;
Kim et al., 2005). Following agricultural fires in
Eastern Europe in spring 2006, Stohl et al. (2007)
measured record high air pollution levels and BC
concentrations in parts of the Arctic above Europe.
Similarly, in a series of studies, Warneke et al. (2009;
2010) found that spring fires in Russia (Siberia) and
Kazakhstan can more than double the Arctic haze
that builds up during the winter months.
2.6.5 BC Impacts in the Himalayas
The world's third largest snowpack after Antarctica
and the Arctic is found in the Hindu Kush-Himalayan-
Tibetan (HKHT) region. The mountain ranges that
define this region fall primarily along the borders of
Pakistan, Afghanistan, India, Nepal, and China (UNEP,
2008a). It is often referred to as the Earth's "third
pole." Atmospheric warming associated with BC is
believed to be a significant factor in the observed
increases in melting rates of glaciers and snowpack
in the HKHT (Barnett et al., 2005; Lau and et al.,
2010; UNEP, 2008a; Thompson, 2010). Ramanathan
and Carmichael (2008) and Ramanathan et al. (2007)
suggested that the advection of air warmed by BC
over the Himalayas has played a role comparable to
that of GHGs in the observed retreat of Himalayan
glaciers. A recent study by Carmichael et al.
(2009) also shows that BC throughout Asia has an
atmospheric warming potential of about 55% of that
attributed to CO2.
High radiative forcing estimates have been
calculated for the Himalayas due to the large
amount of mountain snow and ice cover as well as
the proximity to high emissions of BC from parts
of China and the Indian subcontinent. Flanner et al.
(2007) calculated an average forcing in this region
of +1.5 W rrr2 with short-term forcing of up to 20
W rrr2 in the spring. Translating this to temperature,
Flanner et al. (2009) attributed an increase in the
land-averaged March-May surface temperature in
Eurasia of 0.93°C from BC and organic matter in the
atmosphere and deposited on the snow.
BC can alter snowpack and glacier extent and retreat
through two mechanisms, the first being increasing
and decreasing precipitation as discussed in section
2.6.1.4, and the second being local warming,
especially through deposition, increasing the rate
of melt. Lau et al. (2010) found that heating of the
atmosphere by dust and BC leads to widespread
enhanced warming over the Tibetan Plateau and
accelerated snowmelt in the western Tibetan Plateau
and Himalayas. Menon et al. (2010) show observed
trends in snow cover in the Himalayas, with a
spatially heterogeneous pattern of decreases and
increases of up to 17% from 1990 to 2001, where
the area of decreases is much larger than the area
of increases. Menon et al. (2010) simulated similar
heterogeneous snow cover changes due to aerosol
emissions, showing that the influence of the aerosols
was larger than the influence of changing sea
surface temperatures over that time period. Over
Eurasia, Flanner et al. (2009) conducted a modeling
study that found the combination of strong snow
albedo feedback and large fossil fuel and biofuel
emissions of BC and organic matter from Asia
induce 95% as much springtime snow cover loss as
anthropogenic CO2 alone. The effects on glaciers
are not well quantified, but Xu et al. (2009a) found
evidence that soot deposited on Tibetan glaciers has
been a significant contributing factor to observed
rapid glacier retreat. Changes in the timing and
extent of melting may adversely affect regional
freshwater resources in this region, which relies
heavily on this melt (Carmichael et al., 2009).
Report to Congress on Black Carbon
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Chapter 2
2.6.6 Summary of BC Impacts in Key
Regions
As described in the previous sections, the climate-
related effects of BC can vary considerably across
regions. Table 2-7 provides an overview of the
regional variability in terms of BC's effects on
radiative forcing, temperature, precipitation, and
snow and ice across the United States, Asia, and the
Arctic. In addition, Figures 2-13 and 2-15 are useful
for understanding the regional variability of BC's
radiative forcing effects.
2.7 Metrics for Comparing Black
Carbon Impacts to Impacts of Other
Climate Forcers
In response to Congress's request for an assessment
of potential comparative metrics, this section
summarizes a number of different approaches to
comparing the effects of BC to CO2 and other GHGs,
but cautions that there is no one "best" metric;
rather, the utility of a metric depends on the desired
environmental outcome and policy objective.
Therefore, this section begins by introducing the
concept of using metrics for comparing BC-related
Table 2-7. Climate Effects of BC in the United States, Asia, and the Arctic (Summary).
Effects
Radiative Forcing
Effects
Estimates of direct radiative forcing
of BC over the United States range
from 0.1 to0.7Wm2.
South and East Asia have some of
the world's highest estimates of
radiative forcing, but large ABCs
exert a counterbalancing dimming
effect at the surface.
Average annual snow and ice albedo
forcing in the Tibetan Plateau has
been estimated to be 1.5 W m2, with
local instantaneous forcing up to 20
Wm2.
Arctic
Springtime Arctic forcing has
been estimated to be 1.2 W
m2(direct) and 0.53Wm2
(snow albedo).
Temperature
Effects
No studies were identified for U.S.
temperature effects from BC. All
global modeling studies include the
temperature effects over the U.S.,
but results are difficult to extract.
Estimates of average warming from
BC in the Northern Hemisphere
range from 0.29°C to 0.54°C.
Over the Himalayan region,
atmospheric BC was estimated to
result in up to 0.6°C of warming.
BC deposited on snow results
in warming of roughly 0.4 to
0.5°C, varying by season.
Atmospheric BC was
estimated to contribute
roughly 0.2°C in spring, 0.1°C
in summer, and nearly zero in
autumn and winter.
Precipitation
Effects
One study found little change in
the amount of precipitation in the
western United States as a result of
BC effects.
Other studies have found that
rainfall patterns in the eastern
United States match PM emissions,
but not specifically those of BC.
The cooling at the surface leads
to reduced evaporation and
precipitation as well as changes in
sea-land temperature gradients.
Precipitation and temperature
gradient modifications can lead to
shifts of regional circulation patterns
such as a decrease in the Indian and
Southeast Asian summer monsoon
rainfall and a north-south shift in
eastern China rainfall.
No studies were identified for
Arctic precipitation effects.
Snow and Ice
Effects
In the western United States, BC
deposition on mountain glaciers and
snow produces a positive snow and
ice albedo effect, contributing to the
snowmelt earlier in the spring.
Early snowmelt reduces the amount
of water resources that normally
would be available later in the spring
and summer, and may contribute to
seasonal droughts.
BC atmospheric warming is believed
to be a significant factor in the
melting of the HKHT glaciers and
snowpack.
The deposition of BC on glaciers
and snowpack in Asia also has
a strong snow and ice albedo
positive feedback that accelerates
melting of the glaciers and snow,
with implications for freshwater
availability and seasonal droughts.
BC may increase snowmelt
rates north of 50°N latitude by
as much as 19-28%.
Soot deposition in the Alaskan
Arctic tundra created snow
free conditions five days
earlier than model runs
without BC deposition.
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Black Carbon and Its Effects on Climate
Increasing Certainty that Meeting Target will Reduce Damages
(i.e.. Policy Relevance)
Emissions
Atmospheric
Concentrations
Radiative
Forcing
Climate
Change
Impacts
>.
Damages
Increasing Certainty of Meeting a Target
Figure 2-24. Cause and Effect Chain from Emissions to Climate Change, Impacts, and Damages.
(Adapted from Fuglestvedt et al., 2003.) The arrows indicate that a policy could focus on different
elements along the causal chain and, depending on whether the policy focuses on the emissions or
damages end of the chain, can determine the certainty of meeting the stated policy target versus the
certainty of reducing damages at issue.
impacts to those of other climate forcers. It explains
some of the approaches to developing metrics and
provides a comparison of common metrics used
for GHGs and for BC. This section concludes with a
discussion of the most salient limitations associated
with specific metrics and with using metrics in
general.
The goal of a metric, as used in this report, is to
quantify the impact of a pollutant relative to a
common baseline. Such metrics can be used to
compare between two or more climate forcers (e.g.,
CO2 versus CH4), or to estimate the climate effects of
different emissions sources (or mitigation measures).
Metrics that enable comparisons among pollutants
or sources based on common denominators can also
be used for the implementation of comprehensive
and cost-effective policies in a decentralized manner
(e.g., in a market-based climate program) so that
multi-pollutant emitters can compose mitigation
strategies (Forster et al., 2007).
Climate metrics are often defined relative to a
baseline pollutant (usually CO2) and focus on a
particular climate impact (such as radiative forcing
or temperature) that would be altered due to a
change in emissions. For example, in EPA's annual
Inventory of U.S. Greenhouse Gas Emissions and
Sinks, the GWP metric is used to convert all GHGs
into "CO2-equivalent" units. Importantly, metrics
such as GWP have been used as an exchange rate in
multi-pollutant emissions policies and frameworks
(IPCC, 2009). The key assumption when developing
a metric is that two or more climate forcers are
comparable or exchangeable given the policy
goal. That is to say, one pound of apples may be
comparable to or exchangeable with one pound of
oranges if the goal is not to overload a truck, but
not if the goal is to make apple cider (Fuglestvedt et
al., 2010). Therefore, when used as an exchange rate
in multi-pollutant emissions framework, a metric
allows substitution between climate forcers which
are presumed to be equivalent for the policy goals
(Forster et al., 2007).
Metrics can also be used to prioritize among
mitigation measures designed to control emissions
of similar compounds from different sources. As
described previously in this chapter, aerosols are
composed of numerous components, and these
different components can contribute to both
warming (BC) and cooling. A metric can aggregate
these effects in order to determine the relative
contribution of a given source or measure.
2.7.1 Metrics Along the Cause and Effect
Chain
For both BC and GHGs, there is a cause and effect
chain starting with anthropogenic emissions and
leading to changes in concentrations, radiative
forcing, physical climatic changes, and impacts on
human and natural systems (Figure 2-24). Some
of the links in this cause and effect chain may be
simultaneous rather than sequential. For example,
the atmospheric loading of aerosols affects
dimming and precipitation directly, rather than
mediated through radiatively induced temperature
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changes. Nor is the chain always unidirectional.
Climatic changes can lead to changes in atmospheric
concentrations of climate-forcing pollutants (e.g.,
changes in precipitation will change aerosol lifetimes)
or even emissions of those pollutants (e.g., changes
in temperature affect fossil fuel consumption for
heating and cooling needs, which affects emissions
of particles and precursors). There are uncertainties
at each stage of the cause and effect chain, and
these uncertainties compound over multiple steps
of the chain. The uncertainties for BC are generally
larger at all stages of the causal chain compared to
the long-lived GHGs (for reasons discussed in this
and other chapters of this report).
Within the climate change field, metrics have been
calculated for changes in radiative forcing, global
mean temperature, and monetized damages. The
closer the metric is to the emissions end of the chain,
the less uncertainty there is in how to calculate the
metric; it is easier to determine how a change in
emissions will change concentrations than it is to
determine how a change in emissions will change
temperature (a calculation which requires several
intermediate steps). Additionally, the further along
the chain, the more physical systems (and economic
systems) need to be included in order to calculate
the metric. However, if a reduction in damages is
considered the ultimate objective of the policy,
then a metric that focuses explicitly on impacts or
damages best represents that objective. Since the
economic value of damages (expressed in dollars)
is one of the easiest metrics for the public and
policymakers to place in context, there has been
a great deal of interest recently in calculating the
monetary value of climate change impacts associated
with different pollutants (see Chapter 6). The choice
of a metric can be considered in part a choice about
how to allocate uncertainty between calculation of
the metric and the representativeness of the metric
for the ultimate impacts of interest.
Fuglestvedt (2009) identified the following
considerations for developing a metric for
climate forcers (see Table 2-8 for examples of
how commonly used metrics address these
considerations):
1. What climate impact is of interest for the policy
being considered?
2. What climate forcer will be used as the baseline
for comparison?
3. What is the temporal frame for emissions? Is it
an instantaneous pulse or a sustained change in
emissions?
4. What is the temporal frame for the impact?
10 years, 50 years, 100 years? Is the impact
considered only at the end point of the time
frame, or integrated over the period?
5. Does the metric address the magnitude of
change or the rate of change or both?
6. What is the spatial dimension of the metric
for both emissions and impacts? Is it global or
regional?
7. What economic considerations should be taken
into account? How are damages in the far future
weighed compared to damages in the near term?
Table 2-8. Examples of Commonly Used Metrics for GHGs.
Metric Type Climate Impact Bcaseline Emissions Spatial Includes Rate
'r r Forcer Type Scale of Change?
GWP (Global Warming Potential)
GTP-pulse (Global Temperature Potential)
GTP-sustained
STRE (Surface Temperature Response per
unit continuous Emission)
SFP (Specific Forcing Pulse)
Cost-effectiveness Metrics (e.g., Manne
and Richels, 2001, Global Cost Potential)
Value of Damages (e.g.. Social Cost of
Carbon, Global Damage Potential)
Integrated radiative
forcing
Temperature
Temperature
Temperature
Energy
Mainly temperature
Range of climate
damages
C02
C02
C02
C02
Joules/gram
C02or $ value
$ value
Pulse
Pulse
Sustained
Sustained
Pulse
Optimal
emissions
calculation
Pulse
Global
Global
Global
Global
Global or
regional
Global
Global
No
No
No
No
No
Optional
Limited
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First, the climate impact must be identified because
the effectiveness of a given metric is dependent on
the primary policy goal. Considerations 2 through
7 are then framed by the selected climate impact.
This is important because choosing an inappropriate
metric could lead to policy decisions that ultimately
result in undesirable climate or economic impacts.
2.7.2 Commonly-Used Metrics for GHGs
Article 2 of the United Nations Framework
Convention on Climate Change (UNFCCC) calls for
a policy that addresses the magnitude and the rate
of climate change as well as the cost effectiveness
of controlling emissions (IPCC, 2009). Therefore,
appropriate metrics could cover either the physical
or economic dimensions of climate change, or both.
A number of metrics have been developed and
refined for application to CO2 and other long-lived
GHGs. These metrics are summarized in Table 2-8
and described further below. Their potential
applicability to BC is considered in the next section.
Note that two of the metrics listed in the table (SFP
and STRE) were developed specifically for application
to short-lived climate forcers like BC, and are
discussed only in section 2.7.3.4 below.
Five considerations are listed in Table 2-8. The first,
climate impact, refers to where the metric falls on
the cause-effect chain shown in Figure 2-24. The
second, baseline forcer, lists whether the metric is
measured in comparison to CO2, or in absolute units
(whether dollars or energy). The third column notes
what kind of emissions change is being considered.
A "pulse" of emissions refers to an effectively
instantaneous release of that pollutant (though
sometimes that release is considered to be spread
out over a year). A pulse analysis is appropriate for
a one-time trading of emissions permits, but may
not be as realistic for analyzing investment decisions
which spread reductions out over time (though a
longer term reduction can be approximated as a
series of pulses). Therefore, other analyses consider
the possibility that an emission reduction (or
increase) will be permanent (i.e., sustained overtime).
The third temporal option is to calculate the optimal
emissions path, which is discussed in more detail
in section 2.7.2.3 (cost-effectiveness metrics). The
fourth column shows that most metrics have been
designed to be used on a global scale, though some
of these might be adaptable for regional impacts.
Finally, most metrics consider temperature change or
damages either at a single point in time or summed
over time: only a few consider that there may be
value in limiting the rate of change in addition to
reducing the absolute magnitude of the change.
Table 2-8 is also ordered in a rough approximation
of the transparency of the metric. Metrics which
are transparent and easy to calculate are likely to
be more readily accepted for policy use than those
which require complex modeling. The GWP is in
widespread use and can be calculated based only
on knowing the average lifetime of a molecule
of a gas and the radiative forcing caused by that
molecule. The remaining metrics require the use
of computer models of more or less complexity in
order to calculate, and if the metric is sensitive to
assumptions involved in the modeling then that
could have potential for controversy.
2.7.2.1 Global Warming Potential
To date, the most widely established and well-
defined metric is the GWP. The definition of the GWP
by the IPCC (2007) is
"An index, based upon radiative properties of well-
mixed greenhouse gases, measuring the radiative
forcing of a unit mass of a given well-mixed
greenhouse gas in the present-day atmosphere
integrated over a chosen time horizon, relative
to that of carbon dioxide. The GWP represents
the combined effect of the differing times these
gases remain in the atmosphere and their relative
effectiveness in absorbing outgoing thermal
infrared radiation. The Kyoto Protocol is based on
GWPs from pulse emissions over a 100-year time
frame."
The identified climate impact the GWP addresses is
globally averaged change in radiative forcing and its
baseline climate forcer is CO2 (e.g., the GWP of CO2 is
defined to be I).15 The temporal frame for emissions
is a pulse. The GWP provides the magnitude, but
not the rate of change, of the integrated radiative
forcing over a given time frame. The time frame is
usually 100 years, but 20-year and 500-year GWPs
are sometimes presented to show how GWPs would
differ if short-term or long-term impacts are given
more weight. Finally, the GWP, which addresses only
radiative forcing, a physical metric, does not take
into account any economic dimension.
As discussed below, there have been a number of
criticisms of the GWP in the peer-reviewed literature
(e.g., O'Neill, 2000; Shine, 2009), mainly focused
on either the inability of the GWP to capture key
differences between gases (such as different
lifetimes) or the failure of the GWP to incorporate
economic considerations. Despite such criticisms, at
the time of the Kyoto Protocol in 1997, the GWP was
15 The GWP is calculated as the ratio of the Absolute Global
Warming Potential (AGWP) of a given gas to the AGWP of
CO2. The AGWP has units of W m 2yr g4.
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adopted as the metric used in climate negotiation.
While acknowledging that there are shortcomings
involved in using GWPs even for comparisons among
the long-lived gases, a recent IPCC Expert Meeting
on the topic found that GWPs were still a useful
measure for these gases (IPCC, 2009). It remains
the most accepted metric due to its simplicity, the
small number of input parameters, the relative ease
of the calculation, and a lower level of uncertainty
compared to some alternatives (Shine et al., 2005).
The GWPs as calculated by the IPCC Second
Assessment Report (Schimel et al., 1996) currently
remain the standard GWPs used for the official U.S.
GHG emissions inventory compiled annually by EPA,
as required by UNFCCC reporting guidelines.16
2.7.2.2 Global Temperature Potential
One alternative metric that has received recent
attention is the GTP. Like the GWP, the GTP is
a physical metric. Whereas the GWP considers
change in globally averaged radiative forcing, the
GTP compares the globally averaged temperature
change at a given point in time resulting from the
emission of two climate forcers of equal mass (Shine
et al., 2005). The GTP moves one step further down
the cause and effect chain and addresses a climate
response to radiative forcing, the global-mean
surface temperature change. The GTP therefore
includes more physical processes, such as the heat
exchange between the atmosphere and ocean, than
the GWP. This introduces more uncertainty to the
metric, and can require the use of more complex
models in order to calculate the GTP value. In
addition, while the GWP represents the integrated
radiative forcing of a pulse of emission over a given
time period, the GTP is evaluated at a given point
in time (IPCC, 2009). Like GWPs, the GTP can be
calculated over a variety of timescales, with 20,100,
and 500 years being the timescales most commonly
presented. There are advantages and disadvantages
to using either the GWP or a GTP, and they may each
address different policy goals and may be more
relevant to different climate forcers and time frames,
depending upon the policy need. To date, however,
the GTP has not been used as a metric for trading
gases in international, national, or regional accords.
There are two versions of the GTP: one that involves
the effects of a pulse of emissions, and another
that involves a sustained reduction of emissions.
The latter version of the GTP results in comparative
values between different gases that are similar
to the values calculated using GWPs. The former
version of the GTP, by contrast, leads to longer-lived
16 See http://www.epa.gov/dimatechange/emissions/
usgginventory.html.
gases being given more relative weight because a
pulse of a short-lived gas has very little impact on
temperatures many years in the future.
The GTP can also be calculated as a function of a
future global temperature stabilization target. One
criticism of a number of metrics is that they are not
compatible with a goal of stabilization because the
target is not part of the metric. Manne and Richels
(2001) developed a methodology to calculate a
time-dependent metric (referenced below as a cost-
effectiveness metric) that would change as a target
level was approached. For example, if the target is
not to exceed a 2 degree temperature change above
preindustrial, then when global temperatures are
still only 1 degree above preindustrial, and therefore
the target temperature is still decades away, the
metric will place weight on long-lived gases like
CO2. But as the target temperature is approached,
the time to reach that target becomes short, and
the metric places weight on the strong, short-lived
forcers like CH4 and BC.
Shine et al. (2007) used a similar approach to
develop a time-dependent GTP, the GTP(t). Shine
et al. applied the GTP(t) to BC using a target of 2°C,
and found that for a low emissions scenario, GTP(t)
starts at about 2 in 2010, rising to 1,000 by 2080. But
for a high emissions scenario, GTP(t) can start at 200,
reach 1,000 in 2030, and reach 20,000 in about 2045.
While this approach is one of the few approaches
that are truly compatible with a stabilization target,
there are some drawbacks. Drawbacks include
the dependence on assumptions about future
emissions scenarios, the undefined nature of
the metric after reaching the stabilization target,
and the dependence of the metric on computer
modeling, which reduces transparency. In addition,
policymakers might not desire a metric whose value
can change by orders of magnitude over several
decades and without a transparent and predictable
schedule. One advantage of the GTP(t) and related
metrics is that they can easily be adapted to include
a rate of change goal; for example, rather than just
constraining the metric to reach a 2°C target, it is
also possible to value the rate of change as well
by adding on a constraint that the temperature
not increase more than a given amount in any
given decade. Such an additional constraint would
increase the value of short-lived substances like BC.
2.7.2.3 Cost-Effectiveness Metrics
Manne and Richels (2001) examined relative
tradeoffs between different gases that vary over
time and are calculated to optimally achieve a
given target using a computer model that included
economic considerations. Similarly, the Global Cost
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Potential (GCP), compares the relative marginal
abatement costs for two climate forcers when a
given climate change target is achieved at least cost
(IPCC, 2009). These approaches define a temperature
or radiative forcing target and calculate the relative
(or absolute) dollar value that should be imposed on
different gases in order not to exceed that target.
2.7.2.4 GHG Metrics for Measuring Economic
Impacts
Two metrics, the Global Damage Potential (GDP) and
the social cost of a pollutant, involve monetization
of the damages of climate change (see detailed
discussion in Chapter 6). The GDP compares the
relative damage resulting from an equal mass of
emissions of two climate forcers (IPCC, 2009). The
social cost calculation has most commonly been
used for CO2 alone, where it is referred to as the
Social Cost of Carbon (SCC). However, even where
risks and impacts can be identified and quantified
with physical metrics, it may be difficult to monetize
these risks and impacts (e.g., such as ecosystem
damage or the potential to increase the probability
of an extreme weather event) such that an accurate
cost-benefit comparison could be undertaken. Both
the GDP and the social cost calculation depend on
the physical aspects of the climate system as well
as the economic linkages between climate change
impacts and the economy (IPCC, 2009). Therefore,
the GDP and the social cost require calculations of
the entire cause and effect chain, but as a result
contain a large amount of uncertainty.
2.7.3 Applicability of Climate Metrics to BC
This section discusses the use of well-established
metrics such as the GWP and GTP as they relate to
BC emissions and identifies alternative metrics that
may be more relevant to BC. As discussed earlier
in this chapter, BC influences the climate differently
than the warming effects of GHGs. These differences
have important implications for identifying
appropriate metrics to compare climate impacts (and
reductions thereof). Table 2-1 compared some of
BC's climate attributes and effects to those of CO2.
The implications of these differences with respect to
metrics are discussed here.
As described in detail below, the significant
differences between BC and CO2 make applying the
metrics introduced in the previous section difficult
and, for some purposes, inappropriate. One of the
most essential factors to consider is that BC is most
clearly related to short-term climate impacts, and is
principally a regional pollutant. The lifetime of BC
(weeks) is much shorter than the mixing time of the
atmosphere (1 to 2 years), so the climate impacts of
BC depend heavily on where and when it is emitted.
In comparison, the shortest-lived GHG in the Kyoto
basket has a lifetime longer than one year, and the
majority of the Kyoto gases have lifetimes ranging
from decades to millennia. In addition, the variation
in atmospheric concentrations of BC among regions
contrasts with the well-mixed nature of most GHGs.
This distinction has not been captured in most
metrics to date. Thus, focusing on long-term, global
average radiative forcing impacts— the frame
of reference for long-lived GHGs — may lead to
distorted policy decisions about BC. Conversely,
focusing on short-term or regional impacts may
be inappropriate for decisions involving long-lived
GHGs. The following sections discuss how different
physical (GWP, GTP, SFP, and STRE) and economic
metrics have been used to compare BC to other
substances.
2.7.3.1 Global Warming Potential
While a GWP can be calculated for BC, there are
reasons that GWPs may be less applicable for this
purpose due to the different nature of BC compared
to GHGs, in terms of various physical properties and
the fact that unlike GHGs, BC is not well mixed in the
atmosphere. However, because GWPs are the most
commonly used, and only official, metric in climate
policy discussions, many studies have calculated
GWPs for BC. One-hundred-year GWPs for BC in the
literature range from 330 to 2,240. That is to say, 330
to 2,240 tons of CO2 would be required to produce
the same integrated radiative effect over 100 years
as one ton of BC. Some of the factors that account
for the range in these estimates include the use of
different and uncertain indirect and snow/ice albedo
effects estimates, use of a different CO2 lifetime for
the baseline, and recognition of the dependence of
a GWP for BC on emissions location.
Using time periods shorter than 100 years has also
been explored for determining the GWPs of BC.
Those who are concerned with near-term impacts
(such as Arctic ice retreat) sometimes suggest
20-year GWPs as more appropriate for short-lived
forcers such as BC (CATF, 2009b). Jacobson (2007)
estimates a 20-year GWP for BC of 4,470. However,
for those concerned about the long-term problems
of climate change, even 100-year GWPs may be
considered too short (IPCC, 2009). Because BC is a
short-lived species, the shorter the policy-relevant
time horizon considered, the greater the relative
importance of BC compared to CO2 (and vice
versa: the longer the relevant time horizon, the less
important BC is compared to CO2). If the focus is on
achieving immediate climate benefits within a 10- to
20-year time period, the 20-year GWP provides a
more realistic picture of the impact of reductions
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in different species in the near-term. On the other
hand, if the concern is to identify measures that will
help avert climate change at a broad scale, over
a longer time frame, as the problem is generally
conceptualized, a 20-year time horizon is insufficient,
and the 100-year GWP is a more relevant metric.
2.7.3.2 Global Temperature Potential
GTPs, as described previously, evaluate the impact
on temperature at a given time. Studies have applied
the GTP using approaches that differ with respect to
how the emissions are reduced and how the impacts
are calculated. Boucher and Reddy (2008) use a
short, pulse-like (1-year) reduction of emissions and
find that the 100-year GTPs are about a factor of 7
smaller than the corresponding GWPs. Berntsen et al.
(2006) reduced BC emissions for a 20year time span
(approximately the lifetime of a given investment
in abatement technology) and found that the 100-
year GTP of BC was about 120 to 230 (i.e., reducing
120 to 230 tons of CO2 has the same impact on
temperatures in 100 years as reducing 1 ton of BC).
Several papers have recently summarized different
BC GWP and GTP estimates (Sarofim, 2010; California
Air Resources Board, 2010; Fuglestvedt et al., 2010).
However, of the studies surveyed by these three
papers, only Hansen et al. (2007a) considered
indirect cloud interactions of BC and only a few
included estimates for metrics of co-emitted OC. If
co-emissions are not included, then any metric will
likely overestimate the globally averaged climate
benefits of reducing BC. Inclusion of indirect effects
could either increase or decrease the calculated
value of the metric.
Figure 2-25, based on Fuglestvedt et al. (2010),
summarizes a number of studies that attempted
to develop metrics for comparing CO2 and BC. This
figure shows how the GWP metric depends on the
time horizon used (20 years, 100 years, and 500
years). Additionally, for the first four studies, the
range of values results from a dependence of the
GWP on the region in which the emission occurs.
The difference between the studies is the result of
differences in the climate models used to link the
emissions to the temperature change. Figure 2-26
shows a similar analysis from Fuglestvedt et al.
(2010) which evaluates the equivalent GTP for these
different models.
Fuglestvedt et al. (2010) show that the metric for
comparing BC to CO2 can range from a ton of BC
being equivalent to 48 tons of CO2 based on a
100-year GTP (which measures the temperature
change 100 years after a pulse of emissions) in
one model, to 4,900 tons of CO2 based on a 20-
year GWP (which integrates the total radiative
forcing impact of a pulse of emissions over the
20-year time span) in another model. The variation
between GWPs or GTPs for emissions from different
Koch et al.
(2007b)
Naiketal.
(2007)
20-year Horizon
Reddy and Berntsen et al. Bond and Sun Schulzetal.
Boucher(2007) (2006) (2005) (2006)
100-year Horizon
500-year Horizon
Figure 2-25. Ranges and Point Estimates for Regional Estimates of GWP Values for One-Year Pulse
Emissions of BC for Different Time Horizons. The GWP values in the Y axis of the figure refer to
the number of tons of CO2 emissions which are calculated to be equivalent to one ton of BC
emissions based on the particular metric. (Adapted from Fuglestvedt et al., 2010.) Note that the
first four studies referenced evaluated GWP values for different sets of regions; Bond and Sun
(2005) and Schulz et al. (2006) produced global estimates only.
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Koch et al. Naiketal. Reddyand Berntsen et al. Schulzetal.
(2007b) (2007) Boucher (2007) (2006) (2006)
• 20-year Horizon • 100-year Horizon • 500-year Horizon
Figure 2-26. Ranges and Point Estimates for Regional Estimates of GTP Values for One-Year Pulse
Emissions of BC for Different Time Horizons. The GTP values in the Y axis of the figure refer to
the number of tons of CO2 emissions which are calculated to be equivalent to one ton of BC
emissions based on the particular metric. (Adapted from Fuglestvedt et al., 2010) Note that the
five studies referenced evaluated radiative forcing estimates for different sets of regions (which were
translated into GTP values by Fuglestvedt et al.); Schulz et al. (2006) produced global estimates only.
locations demonstrates how variability in convective
properties, exposure to sunlight, and different
surface albedos can cause the effect of a given unit
of emissions of BC to vary. Given a specific timescale,
metric, and computer model, the two figures show
that this dependence on emissions location can lead
to changes in GWP or GTP by up to factor of three.
Such dependence on emissions location for long-
lived GHGs does not come into play when calculating
their GWPs.
Sarofim (2010) also summarized a number of
studies, and further analyzed how the GWP estimate
depended on inclusion of either fossil fuel OC co-
emissions or snow albedo impacts. Sarofim (2010)
found that inclusion of these processes can change
the value of the metric by about a factor of two.
Other effects that were not quantified in the paper,
but that can lead to significant differences between
model estimates of GWPs, are the inclusion of
indirect effects on clouds and the assessment of
a larger range of sectors and co-emission types.
Additionally, because most metrics use CO2 as a
baseline forcer, the use of different carbon cycle
models can significantly influence the metric values
for BC. Some researchers may report metric values
in carbon equivalents, rather than CO2 equivalents,
which leads to a factor of 3.7 difference.
2.7.3.3 Specific Forcing Pulse
The SFP is a relatively new metric proposed by
Bond et al. (2011) to quantify climate warming or
cooling from short-lived substances (i.e., substances
with lifetimes of less than four months). This metric
is based on the amount of energy added to the
Earth system by a given mass of the pollutant.
The rationale for developing this new metric was
that short-lived substances contribute energy on
timescales that are short compared to time scales of
mitigation efforts, and therefore can be considered
to be "pulses." Bond et al. (2011) find that the SFP
of the direct effect of BC is 1.03±0.52 GJ/g, and with
the snow albedo effect included is 1.15±0.53 GJ/g.
They also find that the SFP for OC is -0.064 (from
-0.02 to -0.13) GJ/g, which leads to a conclusion that
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for direct forcing only, a ratio of about 15:1 for OC
to BC is close to climate neutral. However, this does
not include cloud indirect effects or co-emissions of
substances other than OC. Bond et al. also find that
the SFP varies by 45% depending on where the BC
is emitted. While the paper notes that fundamental
differences in temporal and spatial scales raise
concerns about equating the impacts of GHGs and
short-lived aerosols, they do use the SFP to calculate
a GWP for the direct effect of BC of 740 (±370), for
both the direct and the snow albedo effect of BC of
830 (±440), and for organic matter of -46 (from -18
to -92).
This metric is mathematically similar to the Absolute
Global Warming potential (see footnote in GWP
section), but is applied somewhat differently.
Additionally, the use of this metric for regional
impacts is interesting, though as discussed earlier,
the regional pattern of radiative forcing (or energy
input) is not necessarily the same as the regional
pattern of temperature response to that forcing.
2.7.3.4 Surface Temperature Response per Unit
Continuous Emission
Another new metric, the STRE has been proposed by
Jacobson (2010). The STRE is similar to the sustained
version of the GTP. Jacobson found that the STRE
(which he compares to GWPs) for BC on the 100 year
time scale is 2,900 to 4,600 for BC in fossil fuel soot
and 1,060 to 2,020 for BC in solid-biofuel soot. The
uncertainty ranges presented by Jacobson depend
on his assumption that CO2 will decay exponentially
with either a 30- or a 50-year lifetime. The use of
a more sophisticated carbon cycle model or the
Bern carbon cycle approximation from the IPPC
(which is a sum of 4 exponentials rather than a
single exponential as in the Jacobson calculations)
would result in a lower STRE and would be more
comparable with other approaches. Jacobson also
presents estimates of the combined BC plus OC
STRE, finding that the STRE for emissions of BC plus
OC from fossil fuel soot ranges from 1,200 to 1,900
and for emissions from biofuel soot the STRE ranges
from 190 to 360.
2.7.3.5 Economic Valuation Metrics
Economic valuation approaches for BC that focus
on valuing climate damages from a comprehensive,
societal standpoint are discussed in detail in Chapter
6. For reasons discussed in that chapter, techniques
used to value the climate damages associated with
long-lived GHGs are not directly transferrable to
BC or other short-lived forcers. In fact, most such
approaches have focused exclusively on valuing
the climate impacts of CO2, and may not even be
transferrable to other GHGs. Additional work is
needed to design approaches for valuing the climate
impacts of BC directly, and to incorporate those
approaches into metrics comparable to the SCC.
2.7.4 Using Metrics in the Context of
Climate Policy Decisions
There is currently no single metric widely accepted
by the research and policy community for
comparing BC and long-lived GHGs. In fact, some
question whether and when such comparisons are
useful. For example, there are concerns that some
such comparisons may not capture the different
weights placed on near-term and long-term climate
change. However, there are multiple reasons to
compare BC to other short-lived and long-lived
climate substances, including offsets, credit trading,
evaluation of net effects of a mitigation option, or
illustrative analyses.
The choice of a metric depends greatly on the policy
goal. No single metric will accurately address all
the consequences of emissions of all the different
climate forcers, and all of the differences between
BC and the well-mixed gases must be considered.
The appropriate metric to use depends on a range
of factors, such as: the time scale (20 years, 100
years, or more), the nature of the impact (radiative
forcing, temperature, or more holistic damages),
concern over different processes (indirect effects,
snow albedo changes, co-emissions), and whether
sources and impacts should be calculated regionally
or globally. It is important to note that different
climate models will yield different results even if
the same metric definition is chosen. Taking several
of these factors into account, especially the use of
different time scales, a ton of BC has been calculated
to be equivalent to anywhere from 48 tons of CO2 to
4,600 tons of CO2. For comparison, the UNEP/WMO
assessment, looking only at the 100 year timescale,
estimated that BC could be 100 to 2,000 times as
potent as CO2 per ton (UNEP and WMO, 2011a).
Certainly, the appropriateness of the comparison
depends on the policy question at hand, and the
differences in lifetime, uncertainties, co-emissions,
modes of interaction with the climate system, and
non-climatic effects such as human health should
be evaluated when choosing a metric. This section
highlights how these differences affect the metric
choice.
The tradeoff between capturing short-term
and long-term impacts is not strictly a scientific
consideration but also a policy question. Much
like the original choice of 100 years for the GWP
was a policy compromise between long-term and
short-term impacts; policymakers may consider
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Black Carbon and Its Effects on Climate
whether using a GWP or GTP metric is an acceptable
compromise given a desire to compare BC and the
long-lived GHGs. A key question is how the metric
is used to inform the policy decision. The NRC has
warned against delaying CO2 reductions in favor of
short-lived forcer mitigation, suggesting that CO2
emissions control and control of short-term forcing
agents could be thought of instead as "two separate
control knobs that affect entirely distinct aspects
of the Earth's climate" (National Research Council,
2011). The results of the UNEP/WMO assessment
suggest that the two strategies are complementary
and should be pursued simultaneously, with BC
reductions forming part of a larger strategy for
near-term climate change and CO2 programs
influencing climate over the longer term (UNEP and
WMO, 2011a). Such an approach could incorporate
separate metrics for short-lived and long-lived
species. One metric would be appropriate for
guiding global emissions of climate forcers to
achieve stabilization of GHG concentrations in the
long-term, while another metric would focus on
mitigating near-term warming and could be used to
guide regional emissions reductions in short-lived
climate forcers to reduce the impacts on regional
forcing, precipitation, and ice/snow melt. It is
important to recognize that long-term stabilization
of CO2 concentrations requires limiting total
cumulative emissions of CO2 and that CO2 reductions
today are necessary to achieve climate goals decades
and centuries from now (National Research Council,
2011).
Reductions of BC today do little to achieve climate
goals in the next century: however, they are
important for climate goals in the near future, which
can include reducing impacts on vulnerable regions
such as the Arctic and reducing the rate of near-term
climate change. In addition, if and when we
approach climate stabilization, sustained reductions
in emissions of BC will be important to keep those
peak temperatures lower than they would otherwise
be. Along these lines, the IPCC found that the
complexity of climate change may indicate that a
basket of metrics approach would best capture the
variety of spatial, temporal and uncertain features
(IPCC, 2009). Such a basket approach to addressing
short-lived and long-lived forcers separately (though
not BC specifically) has also been supported by
Jackson (2009) and Daniel et al. (2011).
Outside of the policy context, the use of multiple
metrics can be valuable for illustrative purposes. For
example, Figure 2-26 shows the impact of BC relative
to CO2 on different timescales. Such a figure could
be combined with an analysis such as the Unger et
al. (2010) figure replicated in Figure 2-19 to show the
GTP (or GWP) weighted impact of a set of proposed
mitigation options at 20 years and 100 years (or
some other timescale).
2.7.4.1 Considering the Full Range of BC Effects
As discussed in section 2.6, BC is associated
with complex indirect effects and a number of
hydrological effects that are unrelated to radiative
forcing and that—along with the health effects
discussed in Chapter 3—distinguish it from long-
lived GHGs. These effects include impacts on the
water cycle, inhibition of photosynthesis due to
deposition on plants (Kozlowski and Keller, 1966),
enhancement of soil productivity due to deposition
on soil (Laird, 2008), and effects such as surface
dimming. Capturing these additional effects in
a single global metric is challenging. Even the
current GWP metric continues to see widespread
use despite not capturing the ecosystem effects
of CO2-driven ocean acidification or the health
and agricultural impacts of CH4-induced ozone
production.
For most GHGs, relative radiative forcing is a
reasonable approximation of temperature impacts:
a given W m 2 of CO2 has similar impacts to a
W m 2 of N2O. By contrast, BC forcing includes a
combination of surface dimming and absorption
of both incoming and outgoing radiation at many
wavelengths, while GHGs mainly absorb outgoing
thermal infrared radiation. As discussed in section
2.6.1.4, the temperature change resulting from a
given W m"2 of forcing from the snow albedo effect
may be much greater than the temperature change
resulting from a W m"2 of CO2 forcing, whereas the
result of forcing from BC-related direct effects may
depend on the pattern of BC loading. Inclusion of
the cloud effects of BC makes this metric even more
uncertain.
Further complicating the use of existing metrics
for BC are the significant remaining uncertainties
in estimates of BC forcing, especially regarding the
indirect cloud effects [which can be compared to
the uncertainty in forcing from changes in well-
mixed GHG concentrations, estimated to be only
10% of 2.63 W m-2 (Forster et al., 2007)]. However,
even if BC forcing is at the low end of the range,
a consequence of the globally averaged nature of
common metrics is that the right mix of BC and
OC emissions might have little net global radiative
forcing impact and yet still have significant impacts
on regional precipitation, dimming, and snow melt
as well as possibly on regional patterns of warming
and cooling.
Report to Congress on Black Carbon
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Chapter 2
2.8 Key Gaps in Understanding and
Expressing the Climate Impacts of BC
This chapter has summarized key findings from a
wide range of peer-reviewed studies related to BC
and its effects on climate. The complex atmospheric
chemistry of BC and its regional nature make it a
challenging subject for study. The chapter attempts
to identify where the strength of the evidence
suggests that reasonable conclusions can be drawn
(such as for BC's direct forcing impact, which is
widely understood to lead to warming), and also
highlights those areas where such conclusions
may be premature (such as the net effect of BC,
considering its impacts on clouds and also the
impact of co-emitted pollutants). Despite rapidly
advancing science, there is clearly the need for
additional research, particularly with regard to
BC's effects on clouds and its impacts on radiative
forcing, melting and precipitation in specific regions.
Recent studies have begun to apply more rigorous
modeling and estimation approaches to try to
provide better centralized estimates of BC's direct
forcing impact, its impacts on snow and ice, and its
effects on clouds. Further work is needed to improve
these quantitative estimates and to ensure that the
full range of BC effects on climate is considered.
Key research needs include continued investigation
of basic microphysical and atmospheric processes
affecting BC and other co-pollutants, particularly
with regard to the climate effects of BC-cloud
interactions and aerosol mixing state. In addition,
there is a dearth of research on other types of
light-absorbing carbon, such as BrC, which may
also contribute to climate impacts especially in
sensitive regions such as the Arctic. In general,
further investigation of impacts of aerosols in snow-
and ice-covered regions would be fruitful, along
with additional research on the climate impacts
of emissions mixtures from particular source
categories.
It is also difficult to compare BC directly to CO2 or
other long-lived GHGs. This chapter has explored
some of the metrics that are currently available to
determine how well they perform for purposes of
expressing the climate effects of BC and comparing
BC to CO2. However, there are clear limitations to
using these metrics. In general, there is a strong
need for further refinement of policy-re levant
metrics for BC and other short-lived climate forcers.
Appropriately tailored metrics for BC are needed in
order to quantify and communicate BC's impacts
and properly characterize the costs and benefits of
BC mitigation.
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Chapter 3
Black Carbon Effects on Public
Health and the Environment
3.1 Summary of Key Messages
• BC is a component of both fine and coarse
particulate matter (PM), though because of its
small size, it is most strongly associated with the
fine particle (PM2.5) fraction. Most of the literature
evaluating the potential impacts of BC on human
health (and the health benefits of BC mitigation)
has focused on BC as part of PM2.5.
• Short-term and long-term exposures to PM2.5
are associated with a broad range of adverse
human health effects including respiratory and
cardiovascular effects, as well as premature death.
• Over the past decade, the scientific community
has focused increasingly on trying to identify the
health impacts of particular PM25 constituents,
such as BC. However, EPA has determined that
there is insufficient information at present to
differentiate the health effects of the various
constituents of PM25; thus, EPA assumes that
many constituents are associated with adverse
health impacts.
- The limited scientific evidence that is currently
available about the health effects of BC is
generally consistent with the general PM2.5
health literature, with the most consistent
evidence for cardiovascular effects. However,
study results for BC are variable, and further
research is needed to address remaining
uncertainties.
• PM2.5, both ambient and indoor, is estimated to
result in millions of premature deaths worldwide,
the majority of which occur in developing
countries.
- The World Health Organization (WHO)
estimates that indoor smoke from solid
fuels is among the top ten major risk factors
globally, contributing to approximately 2
million deaths annually. Women and children
are particularly at risk.
- Ambient air pollution is also a significant
health threat: according to the WHO, urban
air pollution is among the top ten risk factors
in medium- and high-income countries.
Urban air pollution is not ranked in the top
ten major risk factors in low-income countries
since other risk factors (e.g. childhood
underweight and unsafe water, sanitation and
hygiene) are so substantial; however, a much
larger portion of the total deaths related to
ambient PM2.5 globally are expected to occur
in developing regions, partly due to the size
of exposed populations in those regions. It
is noteworthy that emissions and ambient
concentrations of directly emitted PM2.5 are
often highest in urban areas, where large
numbers of people live.
• PM25, including BC, is linked to adverse impacts
on ecosystems, to visibility impairment, to
reduced agricultural production in some parts of
the world, and to materials soiling and damage.
3.2 Introduction
This chapter assesses the current scientific
knowledge relating to the public health and non-
climate welfare effects associated with short-term
and long-term exposure to BC. The magnitude
of these impacts in the U.S. and globally is also
addressed.
3.3 Health Effects Associated with BC
3.3.1 Key Health Endpoints Associated with
Exposure to PM
BC is a component of both fine and coarse PM. Since
1997, EPA has recognized the need to regulate fine
and coarse-fraction particles separately. Current
national ambient air quality standards (NAAQS) use
PM2 5 as the indicator for fine particles, and PM10 as
the indicator for thoracic coarse particles. At present,
EPA is undertaking another periodic review of these
standards. As part of this review, EPA has completed
an Integrated Science Assessment for Particulate
Matter (ISA) (U.S. EPA, 2009b) providing a concise
evaluation and integration of the policy-relevant
Report to Congress on Black Carbon
67
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Chapter 3
Table 3-1. Summary of Causal Determinations for Exposure to PM2.5 from 2009 PM ISA. (Source: U.S. EPA)
Exposure Outcome Causality Determination
Short-term exposure to PM25
Long-term exposure to PM25
Cardiovascular Effects
Respiratory Effects
Mortality
Cardiovascular Effects
Respiratory Effects
Mortality
Reproductive and Developmental Effects
Cancer, Mutagenicity, and Genotoxicity
Causal
Likely to be causal
Causal
Causal
Likely to be causal
Causal
Suggestive
Suggestive
science pertaining to the health and environmental
effects of ambient particles. The ISA presents causal
determinations by PM size fraction and exposure
duration (i.e., short-term [days to weeks] or long-
term [months to years]) for the health effects for
which sufficient evidence was available to conclude a
causal, likely to be causal, or suggestive relationship
(Table 3-1). The discussion below is focused on the
health effects with the strongest weight of evidence
(i.e., cardiovascular effects, respiratory effects,
and mortality) and conclusions drawn for these
effects in the ISA. A more limited subset of studies
has evaluated reproductive and developmental
outcomes and cancer effects, but the weight of
evidence for these effects is less substantial.1
A large body of scientific evidence links exposures
to fine particles (i.e., ambient PM2.5 mass
concentrations) to an array of adverse effects,
including premature mortality, increased hospital
admissions and emergency department visits
for cardiovascular and respiratory diseases, and
development of chronic respiratory disease (U.S.
EPA, 2009b). Recent evidence provides a greater
understanding of the underlying mechanisms for
PM2.5-induced cardiovascular and respiratory effects
for both short- and long-term exposures, providing
biological plausibility for the effects observed in
epidemiological studies. This evidence links exposure
to PM2.5 with cardiovascular outcomes that include
the continuum of effects ranging from more subtle
subclinical measures (e.g., changes in blood pressure,
heart rate variability) to premature mortality. These
health effects may occur over the full range of PM2.5
concentrations observed in the long- and short-term
epidemiological studies and EPA has concluded
1 See Sections 7.4 and 7.5 of the PM ISA for an in-depth
characterization of the evidence for an association between
PM2.s and reproductive and developmental effects and
cancer, respectively, (http://cfpub.epa.gov/ncea/cfm/recordisplay.
cfm?deid=216546)
that no discernible threshold for any effects can be
identified based on the currently available evidence.
In reviewing the studies regarding health effects of
PM2.5, EPA has recognized that it is highly plausible
that the chemical composition of PM would be
a better predictor of health effects than particle
size alone (U.S. EPA, 2009b, 6-202). Differences
in ambient concentrations of PM25 constituents
observed in different geographical regions as well as
regional differences in PM25-related health effects
reported in a number of epidemiological studies
are consistent with this hypothesis (U.S. EPA, 2009b,
Section 6.6). Over the past decade, the scientific
community has focused increasingly on trying
to identify the health impacts of particular PM2.5
constituents or groups of constituents associated
with specific source categories of fine particles. The
growing body of evidence for the health impacts
of specific PM2.5 constituents includes evidence of
effects associated with exposure to BC. However, the
ISA concludes that the currently available scientific
information continues to provide evidence that
many different constituents of the fine particle
mixture, as well as groups of constituents associated
with specific source categories of fine particles,
are linked to adverse health effects. While there is
"some evidence for trends and patterns that link
specific PM25 constituents or sources with specific
health outcomes... there is insufficient evidence
to determine if these patterns are consistent or
robust" (U.S. EPA, 2009b, 6-210). Consequently,
research and data collection activities focused
on particle composition could improve our
understanding of the relative toxicity of different
fine particle constituents or groups of constituents
associated with specific sources of fine particles
to inform future regulatory activities and benefits
assessments.
The body of scientific evidence linking exposures
to coarse particles (i.e., ambient PMi0-2.5 mass
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Black Carbon Effects on Public Health and the Environment
concentrations) to health effects is much smaller
than the body of evidence for PM2.5 (U.S. EPA, 2009b).
Similar to PM2.5, the chemical composition of PM10-2.5
can vary considerably by location, but city-specific
speciated PM10-2.5 data are limited. However, PM10-2.5
may contain iron, silica, aluminum, and base cations
from soil, plant and insect fragments, pollen, fungal
spores, bacteria, and viruses, as well as fly ash,
brake lining particles, debris, and automobile tire
fragments. The last four of these components (fly
ash, brake lining particles, debris, and automobile tire
fragments) are associated with urban or industrial
ambient mixes of coarse PM, which are often coated
with BC (See Figure 2-3). Urban or industrial ambient
mixes of coarse PM are dominated by high density
vehicular, industrial, and construction emissions,
and are likely to be associated with adverse health
effects (U.S. EPA, 2006b). While there are no studies
that specifically examine the association between
BC as a component of PMi0-2.5 and health effects,
the current evidence, primarily from epidemiological
studies, indicates that short-term exposure to PMi0-2.5
is associated with effects on both the cardiovascular
and respiratory systems. However, variability in the
chemical and biological composition of PMi0-2.5,
limited evidence regarding effects of the various
components of PM10_2.5, and lack of clearly defined
biological mechanisms for PM10-2.5-related effects are
important sources of uncertainty (U.S. EPA, 2009b).
3.3.2 Health Effects Related to Ambient BC
Concentrations
Some community epidemiological studies have
included BC2 as one of several indicators of fine
particulate air pollution. Of PM2.5 components,
BC is one of the larger contributors to PM2.5 total
mass. For example, Bell et al. (2007) examined
levels of PM components on a national basis, and
identified EC as one of the seven main contributors.
The effects observed with BC in health studies are
similar to those observed for PM2 5 and some other
PM constituents (e.g., nickel, vanadium), suggesting
that these effects are not attributable solely to
BC. Indeed, it would be difficult to separate the
contribution of BC to these associations from that of
co-emitted OC and other correlated and co-emitted
primary pollutants in such studies. Still, these studies
provide generally consistent evidence for an
association between cardiovascular morbidity and BC
concentrations.
2 The monitoring methods used to estimate BC vary, and include
various surrogate measurements such as optical BC and thermal-
optical EC (see Chapter 5 and Appendix 1). Categorization of studies
according to the indicator measurements used should be the focus
of future research.
A number of studies have reported associations
between short-term exposure to BC and
cardiovascular effects (See Table 3-2). Telomere
length attrition, an indication of biological
age that is inversely associated with risk of
cardiovascular disease, was associated with
ambient BC concentrations in the Boston, MA,
area (McCracken et al., 2010). A series of analyses
found that changes in blood pressure (Delfino
et al., 2010; Mordukhovich et al., 2009; Wilker et
al., 2010) and heart rate variability (HRV) (Adar
et al., 2007; Chuang et al., 2008; Gold et al., 2005;
Huang et al., 2003; Park et al., 2005; Schwartz et
al., 2005) were associated with increases in mean
ambient BC concentration. The ST-segment of an
electrocardiograph represents the period of slow
repolarization of the ventricles and ST-segment
depression can be associated with adverse cardiac
outcomes, including ischemia. Delfino et al. (2011)
found positive associations between ST-segment
depression and BC concentrations. Homocysteine,
a sulfur-containing amino acid formed during
metabolism of methionine, is a risk factor for
atherosclerosis, myocardial infarction (MI), stroke,
and thrombosis. Similarly, lower blood DMA
methylation content is found in processes related
to cardiovascular outcomes, such as oxidative stress
and atherosclerosis. Several studies observed an
association between ambient BC concentration
and elevated plasma total homocysteine (Park
et al., 2008; Ren et al., 2010). An additional study
(Baccarelli et al., 2009) observed an association
between lower blood DMA methylation content
and BC concentrations. Cardiac arrhythmia (a
broad group of conditions where there is irregular
electrical activity in the heart) was associated with
increased concentrations of BC in studies conducted
in Boston (Rich et al., 2005; Rich et al., 2006;
Zanobetti et al., 2009; Baja et al., 2010; Dockery et
al., 2005), but not in Vancouver, Canada (Rich et
al., 2004). Another series of analyses has reported
inconsistent associations between BC and blood
markers of coagulation and inflammation, with
some studies finding an effect (Dubowsky et al.,
2006; Ruckerl et al., 2006; Delfino et al., 2009; 2008;
O'Neill et al., 2007), and others finding no effect for
a blood marker with large intra-individual variability
(i.e., B-type natriuretic peptide or BMP) (Wellenius
et al., 2007) or no effects for acute lag periods (i.e.,
48 hours or 1 week) (Zeka et al., 2006). Ambient
concentrations of BC (Peters et al., 2001; Zanobetti
and Schwartz, 2006) and EC (Bell et al., 2009; Peng et
al., 2009; Sarnat et al., 2008; Tolbert et al., 2007; Ito
et al., 2011) were also found to be associated with
hospital admissions and emergency department
visits for cardiovascular outcomes.
Report to Congress on Black Carbon
69
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SI
o
Table 3-2. Summary of Epidemiological Studies of BC and Cardiovascular Health Outcomes. The study reference numbers listed in the
left-hand column are for purposes of cross-reference with Figure 3-1. (Source: U.S. EPA)
n
is-
a
M LenCe c* j i Health . . . . Representative Exposure Selected Effect Estimates
Figures-" °UtCOme Concentration (ug/m') Assessment" (95%CI)b
Short-term Exposure Studies
—
1.
2.
3.
4.
5.
6.
McCracken et
al. (2010)
Delfinoetal.
(2010)
Mordukhovich
etal. (2009)
Wilkeretal.
(2010)
Adar etal.
(2007)
Chuang etal.
(2008)
Gold etal.
(2005)
Boston, MA
Los Angeles,
CA
Boston, MA
Boston, MA
St. Louis, MO
Boston, MA
Boston, MA
Telomere Length
Blood Pressure
Blood Pressure
Blood Pressure
HRV
HRV
HRV
BC
BC
BC
BC
BC
BC
BC
Mean (annual avg): 0.32
Mean (24-havg): 1.67
Mean (7-day moving average):
1.10
Mean (7-day moving average):
0.98
Median (5-min avg periods):
0.285-2.911 (range across
microenvironments)
Median (24- h avg): 0.79
Median (12-h avg): 1.14
Annual outdoor home
concentration estimates
from spatiotemporal
model
Hourly outdoor
home air-pollutant
concentrations
Continuous
measurements from
single monitor averaged
by hour before BP
measurement
Continuous
measurements from
single monitor averaged
by hour before BP
measurement
Two portable
carts containing
continuous sampling
instrumentation
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
Association with leukocyte telomere
length:
-7.6% (-12.8%, -2.1%)
Change in BP:
SBP: 0.22 (-0.65, 1.09)
DBP: 0.36 (-0.1 1,0.83)
Change in BP:
SBP: 1.46 (0.1 0,2.82)
DBP: 0.87 (0.15, 1.59)
Change in BP:
SBP: 3.52 (2.77, 4.26)
DBP: 2.72 (2.31, 3.12)
Percent change in HRV:
SDNN:-5.3(-6.5,-4.1)
RMSSD:-10.7(-11.9,-9.5)
PNN50+1: -13.2 (-15.0, -11.4)
LF: -11. 3 (-13.7, -8.8)
HF: -18.8 (-21.1, -16.5)
LF/HF: 9.3 (7.2, 11.4)
HR: 1.0 (0.8, 1.3)
Relative risk for ST-segment
Depression <1 mm:
1.50(1.19-1.89)
Relative risk for ST-segment
Depression <0.5 mm:
First rest: 3.8 (0.7, 21. 3)
Blood pressure: 5.7 (0.6, 56.3)
Standing: 8.3 (0.8, 81 .9)
Exercise: 0.6 (0.1, 3.1)
Second rest: 2.8 (0.5, 14.3)
Paced breathing: 3.5 (0.5, 23.6)
-------
M LenCe c* j i Health . . . . Representative Exposure Selected Effect Estimates
Figures-" °UtCOme Concentration (ug/m') Assessment" (95%CI)b
7.
8.
9.
10.
11.
—
12.
13.
14.
15.
Parketal.
(2005)
Schwa rtzetal.
(2005)
Delfinoetal.
(2011)
Parketal.
(2008)
Renetal.
(2010)
Baccarelli etal.
(2009)
Baja etal.
(2010)
Dockery etal.
(2005)
Rich etal.
(2005)
Rich etal.
(2006)
Boston, MA
Boston, MA
Los Angeles,
CA
Boston, MA
Boston, MA
Boston, MA
Boston, MA
Boston, MA
Boston, MA
Boston, MA
HRV
HRV
HRV
Plasma Total
Homocysteine
Plasma Total
Homocysteine
Blood DNA
Methylation
Ventricular
Repolarization
Arrhythmia
Arrhythmia
Arrhythmia
BC
BC
BC
BC
BC
BC
BC
BC
BC
BC
Mean (24-h avg): 0.92
Mean (24-h avg): 1.2
Mean (24-h avg): 1.67
Mean (24-h avg): 0.99
Mean (7-day moving average):
0.99
Mean (24-h avg): 0.89
Mean (1-havg): 1.08
Median (2-day avg): 0.98
Median (24-h avg): 0.94
Median (24-h avg): 0.94
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
Hourly outdoor
home air-pollutant
concentrations
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
Percent change in HRV:
SDNN: -3.4 (-10.2, 3.9)
HF: -13.8 (-28.9, 4.4)
LF: -2.4 (-16.2, 13.6)
LF/HF: 13.2 (-1.1, 29.6)
Percent change in HRV:
SDNN: -5.1 (-1.5, -8.6)
r-MSSD:-10.1 (-2.4, -17.2)
PNN50: -16.9 (-6.0, -26.6)
LF/HF: 7.2 (0.7-14.1)
Odds ratio for ST-segment
Depression <1 mm:
2.07(1.30,3.29)
Percent change in total
homocysteine:
3.13(0.76,5.55)
Percent change in total
homocysteine:
0.68 (-0.46, 1.81)
Coefficient for effect on methylation:
LINE-1:-0.09 (-0.15, -0.02)
Alu: -0.02 (-0.08, 0.05)
Change in mean QTc (msec):
1.89 (-0.16, 3.93)
Association with ventricular
arrhythmias:
Recent arrhythmia > 3 days: 1.02
(0.83,1.24)
Recent arrhythmia <3 days: 1.74
(1.28,2.37)
Odds ratio for ventricular
arrhythmias:
0.93(0.74,1.18)
Odds ratio for paroxysmal atrial
fibrillation:
1.46(0.67,3.17)
co
5"
n
a
c-
o
3
o
3
C
c-
o'
a:
a
Er
a
3
a.
o
3
3
-------
SI
Sj
M LenCe c* j i Health . . . . Representative Exposure Selected Effect Estimates
Figures-" °UtCOme Concentration (ug/m') Assessment- (95%CI)b
16.
17.
18.
19.
20.
21.
22.
23.
Zanobetti et
al. (2009)
Richetal.
(2004)
Sorensen etal.
(2003)
Dubowsky et
al. (2006)
Ruckerl etal.
(2006)
Delfinoetal.
(2008)
Delfinoetal.
(2009)
O'Neill etal.
(2007)
Boston, MA
Vancouver,
Canada
Copenhagen,
Denmark
St. Louis, MO
Erfurt,
Germany
Los Angeles,
CA
Los Angeles,
CA
Boston, MA"
Arrhythmia
Arrhythmia
Blood Markers of
Coagulation and
Inflammation
Blood Markers of
Coagulation and
Inflammation
Blood Markers of
Coagulation and
Inflammation
Blood Markers of
Coagulation and
Inflammation
Blood Markers of
Coagulation and
Inflammation
Blood Markers of
Coagulation and
Inflammation
BC
EC
Carbon
Black
BC
EC
BC
BC
BC
Median (6-h avg):
Ambient: 0.72
Indoor: 0.41
Outdoor: 0.50
Mean (24-h avg): 0.8
Median (24- h avg): 8.1 (10-
6/m)
Mean (24-h avg): 0.9
Mean (24-h avg): 2..6
Mean (24-h avg): 2.00
Mean (24-h avg): 1.59 -1.76
Mean (24-h avg): 1.1
Continuous ambient
measurements from
single monitor; indoor
and outdoor measured
continuously at
participants' homes
Continuous
measurements from
single monitor
Personal exposure
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
Outdoor home
measurements
Hourly outdoor home air
pollutants
Continuous
measurements from
single monitor
Odds ratio for maximum T-wave
alternans < 26 uV:
1.42(1.19,1.69)
Odds ratio for defibrillator discharge:
1.06(0.87,1.33)*
*estimated from graph
Association with plasma proteins:
4.1% increase in plasma proteins per
1 x lO'Vm increase in personal CB
exposure
Association with markers of
inflammation:
IL-6:-0.8(-8.9,8.0)
CRP:13(-0.34,28)
WBC: 1.3 (-2.1, 4.8)
Odds ratio for increase in blood
marker above 90* percentile:
CRP: 1.3 (0.7, 2.4)
ICAM-1:2.6(1.7,3.8)
Coefficient for association:
CRP (ng/ml): 585.61
IL-6 (pg/ml): 0.48
sTNF-RII(pg/ml): 135.15
sP-selectin (ng/ml): 1.99
Cu,Zn-SOD(U/gHb): -187.95
Coefficient for association CRP (ng/
ml): 252 (-54, 558)
IL-6 (pg/ml): 0.16 (0.01, 0.31)
sTNF-RII(pg/ml):38(-26, 102)
sP-selectin (ng/ml): 1.19 (-0.52, 2.90)
Cu, Zn-SOD (U/g Hb): -114 (-229, 1)
TNF-a (pg/ml): 0.02 (-0.06, 0.10)
GPx-1 (U/g Hb): -0.47 (-0.97, 0.03)
Percent change in inflammatory
marker:
ICAM-1 (ng/ml): 5.84 (0.87, 11.05)
VCAM-1 (ng/ml): 9.26 (2.98, 15.91)
vWF (proportion): 7.96 (-4.34, 21.84)
n
3-
Q
-------
M LenCe c* j i Health . . . . Representative Exposure Selected Effect Estimates
Figures-" °UtCOme Concentration (ug/m') Assessment" (95%CI)b
24.
25.
26.
27.
28.
29.
30.
31.
32.
Wellenius etal.
(2007)
Zeka etal.
(2006)
Peters etal.
(2001)
Zanobetti
and Schwartz
(2006)
Bell etal.
(2009)
Peng etal.
(2009)
Sarnatetal.
(2008)
Tolbertetal.
(2007)
Ito etal. (2011)
Boston, MA
Boston, MA
Boston, MA
Boston, MA
106 U.S.
Counties
119 U.S.
Counties
Atlanta, G A
Atlanta, G A
New York, NY
Blood Markers of
Coagulation and
Inflammation
Blood Markers of
Coagulation and
Inflammation
ED visits
and hospital
admissions for
CVD
ED visits
and hospital
admissions for
CVD
ED visits
and hospital
admissions for
CVD
ED visits
and hospital
admissions for
CVD
ED visits
and hospital
admissions for
CVD
ED visits
and hospital
admissions for
CVD
ED visits
and hospital
admissions for
CVD
BC
BC
BC
BC
EC
EC
EC
EC
EC
Mean (24-h avg): 0.73
Mean (2-day avg): 0.77
Mean (24-h avg): 1.35
Median (24-h avg): 1.15
Mean (24-h avg): 0.72
Median (24-h avg): 0.58
Mean (24-h avg): 1.4-1.7
Mean (24-h avg): 1.6
Mean (24-h avg): 1.13
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
County-wide averages
County-wide averages
Positive matrix
factorization applied
to measurements from
single monitor
Continuous
measurements from
single monitor
Average of continuous
measurements from
three monitors
"No significant associations
observed between [BC] and BNP
levels at any of the lags examined."
Percent increase in inflammatory
marker:
Fibrinogen: 0.84 (-0.63, 2.31)
CRP:4.51 (-2.03,11.06)
Sediment rate: -4.56 (-25.55, 16.43)
WBC count: -0.63 (-2.45, 1.19)
Odds ratio for Ml:
1.21 (0.87,1.70)
Percent change in Ml admissions:
8.34(0.21,15.82)
Percent increase in health effect
estimate:
25.8 (4.4, 47.2)
Percent increase in CVD hospital
admissions:
0.72(0.43,1.01)
Relative risk of ED visit for CVD:
1.025(1.014,1.036)
Relative risk of ED visit for CVD:
1.015(1.005,1.025)
Percent excess risk:
1.4(0.1,2.7)
o
3
3
-------
n
o
3
O
3
2
S"
n
>r
n
Q
5-
o
3
M LenCe c* j i Health . . . . Representative Exposure Selected Effect Estimates
Figures-" °UtCOme Concentration (ug/m') Assessment" (95%CI)b
Long-term exposure studies
33.
34.
35.
36.
O'Neill etal.
(2005)
Madriganoet
al. (2010)
Alexeeff etal.
(2011)
Can etal. (2011)
Boston, MA
Boston, MA
Boston, MA
Vancouver, BC
Endothelial
dysfunction
Endothelial
dysfunction
Markers of
inflammation
and endothelial
response
CHD
hospitalizations
BC
BC
BC
BC
Mean (24-h avg): 1.0
Mean (24-h avg): 0.84
Mean (24-h avg): 0.42
Mean (24-h avg): 1.19
Continuous
measurements from
single monitor
Continuous
measurements from
single monitor
Validated spatio-
temporal (land use
regression) model
High resolution land use
regression model
Percent change in vascular reactivity
among those with (a) diabetes and
(b)all subjects:
Flow-mediated dilation: (a) -12.6
(-21. 7, -2.4); (b) -9.3 (-17.8, 0.2)
Nitroglycerin-mediated dilation: -(a)
6.6 (-14.0, 1.5); (b) -5.4 (-12.0, 1.7)
Percent change in blood markers:
sVCAM-1: 4.52 (1.09, 7.96)
slCAM-1: -1.37 (-3.89, 1.15)
Percent change in blood markers:
4 wk exposure, sVCAM-1: 1.00 (-0.65,
2.67)
4 wk exposure, slCAM-1: 1.50(0.22,
2.80)
8 wk exposure, sVCAM-1: 1.20 (-0.58.
3.02)
8 wk exposure, slCAM-1: 1.58(0.18,
3.00)
12 wk exposure, sVCAM-1: 1.26
(-0.58,3.14)
12wk exposure, slCAM-1: 1.49(0.04,
2.95)
Relative Risk of CHD hospitalization:
1.01 (1.00,1.03)
n
IS-
Q
i A more complete description or evaluation of the BC monitoring method is beyond the scope of this summary, and may include surrogate measurements.
> SBP=systolic blood pressure; DBP=diastolic blood pressure; SDNN = standard deviation of normal intervals measured between consecutive sinus beats; RMSSD = root mean square
successive difference, a measure of heart period variability; PNN50=number of times per hour in which the change in consecutive normal sinus intervals exceeds 50 milliseconds; LF=low
frequency component of heart rate variability; HF=high frequency component of heart rate variability; HR=heart rate; Line-l=long interspersed nuclear element-1; Alu=short stretch of
DNA; IL-6 = interleukin 6; CRP=c-reactive protein; WBC=white blood cells; ICAM = intercellular adhesion molecule; sTNF-RII=extracellular domain of the tumor necrosis factor receptor;
sp-Selectin = plasma soluble P-selectin; Cu, Zn-SOD=copper/zinc superoxide dismutase; TNF-a=tumor necrosis factor-alpha; GPx-l=glutathione peroxidase; VCAM=vascular cell adhesion
molecule; vWF=vonWillebrand Factor; BNP=B-type natriuretic peptide; MI=myocardial infarction; CVD=cardiovascular disease; ED=emergency department; CHD = coronary heart
disease.
-------
Black Carbon Effects on Public Health and the Environment
The most noteworthy new cardiovascular-related
revelation in recent years with regard to long-term
PM exposure is that the systemic vasculature may
be a target organ (U.S. EPA, 2009b). Endothelial
dysfunction is a factor in many diseases and may
contribute to the origin and/or exacerbation of MI
or ischemic heart disease, as well as hypertension.
Endothelial dysfunction is also a characteristic
feature of early and advanced atherosclerosis.
New evidence supports an association of ambient
BC with decrements in the systemic vasculature.
O'Neill et al. (2005) reported that increases in mean
BC concentration were associated with decreased
vascular reactivity among diabetics, but not among
subjects at risk for diabetes. Several recent studies
(Madrigano et al., 2010; Alexeeff et al., 2011;
Gan et al., 2011) observed that ambient BC was
associated with a marker of endothelial function and
inflammation, and that genes related to oxidative
defense might modify this association. Consistent
with these findings, animal toxicological studies
have shown that BC can affect heart rate variability
(Tankersley et al., 2007; 2004), cardiac contractility
(Tankersley et al., 2008) and oxidative stress response
(Tankersley et al., 2008), providing biological
plausibility for a long-term effect on cardiovascular
health.
Overall, the limited body of evidence suggests that
ambient BC may be associated with a continuum
of effects ranging from more subtle subclinical
measures (e.g. changes in blood pressure, heart
rate variability) to emergency department visits and
hospital admissions for cardiovascular outcomes
(Figure 3-1). Generally, this is consistent with the
association observed for PM2.5 and cardiovascular
outcomes (Janssen et al., 2011), as described above
(Section 3.3.1).
Fewer studies have examined the effects of BC with
respiratory effects (Table 3-3). Clark et al. (2010)
investigated the effect of exposure to ambient air
pollution in utero and during the first year of life
on risk of subsequent asthma diagnosis (incident
asthma diagnosis up to age 3-4) and reported
that BC exposure was associated with a 14% (1-
29%) increase in asthma risk. Delfino et al. (2006)
found associations between airway inflammation
and ambient EC concentrations among asthmatic
children, while Jansen et al. (2005) reported
an association with a marker of pulmonary
inflammation and BC concentrations among older
adults. These results are supported by toxicological
studies reporting evidence of airway inflammation
(Godleski et al., 2002; Saldiva et al., 2002). There
is consistent evidence from a number of studies
that report associations of respiratory symptoms
among both asthmatic and non-asthmatic children
and ambient BC or EC (Kim et al., 2004; Mann et
al., 2010; McConnell et al., 2003; Patel et al., 2010;
Spira-Cohen et al., 2011). Additionally, Suglia et al.
(2008) reported that ambient BC was associated
with decreased lung function among urban women.
Recent studies evaluated the effect of ambient BC
or EC on respiratory hospital admissions and found
statistically significant associations between the
county-average ambient concentrations of BC or
EC and respiratory hospital admissions (Zanobetti
and Schwartz, 2006; Bell et al., 2009; Ostro et al.,
2009). However other studies found less consistent
evidence (Peng et al., 2009; Mohr et al., 2008) or no
evidence (Sarnat et al., 2008; Tolbert et al., 2007) for
an association between ambient EC and respiratory
emergency department visits. Overall, there is
inconsistent evidence for an association between
ambient BC concentrations and respiratory effects.
Similar to what was observed in studies of PM2.5,
studies examining ambient BC report increased
respiratory symptoms in asthmatic children,
but less consistent evidence for an association
with emergency department visits and hospital
admissions.
Several recent epidemiological studies have
examined the association between mortality and
short-term ambient exposure to components of
PM2.5, including BC or EC (Table 3-4). Lippmann
et al. (2006) reported that nickel, vanadium, and
EC were the best predictors, respectively, of PMi0
risk estimates for mortality. Cakmak et al. (2011)
reported an association between increased exposure
to concentrations of EC and increases in all cause
mortality, while Ito et al. (2011) and Ostro et al.
(2007) found positive associations between EC and
cardiovascular mortality. These associations (Ostro
et al., 2007) were higher in individuals with lower
educational attainment and of Hispanic ethnicity
(Ostro et al., 2008). Studies of long-term exposure
to EC (Lipfert et al., 2006; 2009) and BC (Gan et
al., 2011) also report associations with mortality.
Overall, the limited body of evidence examining
the association of ambient BC with mortality has
reported associations with mortality, especially
cardiovascular mortality. This association is
consistent with the evidence for a causal relationship
between PM2.5 and mortality.
3.3.2.1 Health Effects Related to Indicators of
Ambient BC Concentrations
Concentrations of many traffic-generated air
pollutants are elevated for up to 300-500 meters
downwind of roads with high traffic volumes (Zhou
and Levy, 2007). Numerous sources on roads
contribute to elevated roadside concentrations,
including exhaust and evaporative emissions,
Report to Congress on Black Carbon
75
-------
Chapter 3
Series of short- and long-term
exposure studies in various
geographic locations
(Studies #37-40)
Series of short- and long-term
exposure studies in various
geographic locations
(Studies #26-32,36)
Series of short-term
exposure studies in
Boston, MA
(Studies #13-17)
Ventricular
Repolarization
Cardiovascular
Mortality
Hospital Admissions
Changes in
Blood
Pressure
Heart Rate
Variability
Change in
Plasma Total
Homocysteine
Endothelial
Dysfunction
Blood
Markers of
Inflammation
•
Single short-term
exposure study
conducted in
Boston, MA
(Study #12)
•
Series of short-
term exposure
studies in
various
geographic
locations
(Studies #1-3)
•
Series of short-
term exposure
studies in
various
geographic
locations
(Studies #4-9)
•
Series of short-
term exposure
studies in
Boston, MA
(Studies #10-11)
•
H
Series of long-
term exposure
studies in
Boston, MA that
demonstrate
modification by
genetic variants
(Studies #33-35)
T
Series of short-term
exposure studies in
various geographic
locations
(Studies #18-25,35)
Figure 3-1. Conceptual Diagram of the Epidemiological Evidence for the Association of BC with the
Continuum of Cardiovascular Effects, including sub-clinical effects (bottom level of the pyramid) and
clinical effects, increasing in severity moving up the pyramid. It is important to note that the body of
evidence describing the association between BC and cardiovascular effects is much smaller and less
consistent than the one characterizing PM2.5 and cardiovascular effects. The study reference numbers
listed in parentheses correspond to the reference numbers assigned to individual studies in the left-hand
column of Table 3-2 and Table 3-4. For study-specific details, please see Table 3-2 and Table 3-4. (Source:
U.S. EPA)
and resuspension of road dust and tire and
brake wear. Concentrations of several criteria and
hazardous air pollutants are elevated near major
roads. Furthermore, different semi-volatile organic
compounds and chemical components of PM,
including BC, organic material, and trace metals,
have been reported at higher concentrations near
major roads. While this document is focused on the
health effects associated with BC specifically, this
section discusses the mixture of different pollutants
near major roadways, of which BC is a component.
As such, this section emphasizes traffic-related air
pollution, in general, as the relevant indicator of
exposure to BC.
76
Report to Congress on Black Carbon
-------
Table 3-3. Summary of Epidemiological Studies of BC and Respiratory Health Outcomes. (Source: U.S. EPA)
Study
Location
Health Outcome
Representative
Metric Concentration Exposure Assessmenta Selected Effect Estimates (95% Cl)
(ug/m3)
Clark etal. (2010)
Delfinoetal. (2006)
Jansen etal. (2005)
Kim etal. (2004)
Mann etal. (2010)
McConnell etal.
(2003)
Patel etal. (2010)
Spira-Cohen etal.
(2011)
British Columbia
Los Angeles, CA
Seattle, WA
San Francisco, CA
Fresno, CA
12 Southern
California
communities
New York City, NY
South Bronx, NY
Development of
childhood asthma
Exhaled nitric oxide
(biomarker of airway
inflammation)
Exhaled nitric oxide
(biomarker of airway
inflammation)
Bronchitis symptoms
and asthma
Wheeze among
asthmatic children
Bronchitic symptoms
among asthmatic
children
Respiratory
symptoms among
asthmatics and non-
asthmatics
Lung function and
respiratory symptoms
among asthmatic
children
BC
EC
BC
BC
EC
(estimated
from BC)
EC
BC
EC
Mean (24-h avg):
1.34
Mean (24-h avg):
0.71-1.61
Mean (24-h avg):
Central site: 2.01
Indoor: 1.34
Personal: 1.64
Mean (24-h avg):
0.8
Median (24-h avg):
1.3
Mean (24-h avg):
0.71
Median (24-h avg):
0.49-2.4
Mean (24-h avg):
1.9
Land Use Regression (LUR)
Model
Personal exposure and
continuous measurements
from central site monitors
Personal exposures, indoor
monitors, and central
outdoor monitoring site
BC measured at 10 school
sites
Continuous measurements
from single monitor
Annual averages computed
from 2-week averages
measured in each
community
BC measured at 4 high
schools
Personal and outdoor
school site monitoring at 4
elementary schools
Asthma risk due to average exposures
In utero exposure: 1.08 (1.02, 1.15)
First-year exposure: 1.14(1.01, 1.29)
Association between EC and exhaled NO:
Personal exposure: 0.72 (0.32, 1.12)
Central site exposure: 1.38 (0.15, 2.61)
Association between BCand exhaled NO
among asthmatics:
Central site exposure: 2.3 (1.1, 3.6)
Indoor Exposure: 4.0 (2.0, 5.9)
Personal exposure: 1.2 (0.2, 2.2)
Odds ratios of respiratory illness by school-
based BC concentration:
Bronchitis: 1.04 (1.00, 1.08)
Asthma: 1.07 (0.98, 1.17)
Odds ratio of EC and wheeze:
1.12(0.97,1.30)
Bronchitic symptoms as a function of the
4-yr avg EC concentration:
1.64(1.06,2.54)
Odds ratios for respiratory symptoms and
use of asthma medication:
Wheeze: 1.11 (1.00,1.22)
Cough: 0.95 (0.87, 1.03)
Shortness of Breath: 1.26 (1.14, 1.38)
Chest Tightness: 1.11 (1.01,1.24)
Use of Asthma Medication: 1.09 (0.89, 1.33)
Effect estimates for lung function
decrements:
Personal EC, PEF: -9.13 (-19.13, 0.86)
Personal EC, FEV1: -0.02 (-0.09, 0.04)
School-site EC, PEF: -4.58 (-14.01, 4.85)
School-site EC, FEV1: 0.01 (-0.04, 0.07)
Relative Risks for respiratory symptoms:
Personal EC, cough: 1.61 (1.17, 2.21)
Personal EC, wheeze: 1.67 (1.05, 2.66)
Personal EC, shortness of breath: 1.41 (1.01,
1.99)
Personal EC, total symptoms: 1.54(1.13,
2.10)
co
5"
n
n
a
c-
o
3
o
3
C
c-
o'
a:
a
Er
a
3
a.
o
3
3
-------
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Q
5-
o
3
Sugliaetal. (2008)
Bell etal. (2009)
Ostroetal. (2009)
Peng etal. (2009)
Mohr etal. (2008)
Sarnat etal. (2008)
Tolbert etal. (2007)
Zanobetti and
Schwartz (2006)
Boston, MA
106 U.S. Counties
6 CA Counties
119 U.S. Counties
St. Louis, MO
Atlanta, GA
Atlanta, GA
Boston, MA
Lung function among
women
ED visits and hospital
admissions for
respiratory effects
ED visits and hospital
admissions for
respiratory effects
ED visits and hospital
admissions for
respiratory effects
ED visits and hospital
admissions for
respiratory effects
ED visits and hospital
admissions for
respiratory effects
ED visits and hospital
admissions for
respiratory effects
ED visits and hospital
admissions for
respiratory effects
BC
EC
EC
EC
EC
EC
EC
BC
Mean (predicted
annual): 0.62
Mean (24-h avg):
0.72
Mean (24-h avg):
1.0
Median (24-h avg):
0.58
Median (24-h avg):
-0.6-0.7
Mean (24-h avg):
1.4-1.7
Mean (24-h avg):
1.6
Median (24-h avg):
1.15
Local BC levels estimated
by spatiotemporal land-use
regression model
County-wide averages
County-wide averages
County-wide averages
Continuous measurements
from single monitor
Positive matrix
factorization applied to
measurements from single
monitor
Continuous measurements
from single monitor
Continuous measurements
from single monitor
Effect estimates for change in lung
function:
FEV,: -1.09 (-2.5, 0.3)
FVC: -0.62 (-1.9, 0.6)
FEF2sW -3.03 (-5.8, -0.3)
Percent increase in health effect estimate:
511 (80.7,941)
Excess risk per IQR:
Ages <1 9: 5.4 (0.8, 10.3)
Ages <19, Cool season only: 6.8 (-0.2, 14.2)
Percent increase in respiratory hospital
admissions:
0.43 (-0.02, 0.85)
Relative risk for asthma ED visits:
Ages 2-5, summer: 1.01 (0.93, 1.09)
Ages 6-10, summer: 0.98 (0.77, 1.26)
Ages 11-17, summer: 1.09 (1.02, 1.17)
Ages 2-17, summer: 1.05 (1.00, 1.11)
Relative risk of ED visit for Resp:
0.996(0.988,1.003)
Relative risk of ED visit for respiratory
effects:
0.996(0.989,1.004)
Percent change in pneumonia ED visits:
11.71 (4.79,17.36)
i A more complete description or evaluation of the BC monitoring method is beyond the scope of this summary, and may include surrogate measurements.
> NO = nitric oxide; PEF=peak expiratory flow; FE\A=forced expiratory volume in 1 second; FVC=forced vital capacity; FEF2575%=difference between the 25th and 75th percentile of forced
expiratory flow; ED=emergency department.
-------
Table 3-4. Summary of Epidemiological Studies of BC and Mortality. The study reference numbers listed in the left-hand column are for
purposes of cross-reference with Figure 3-1. (Source: U.S. EPA)
Reference H itu Representative _
Number in Study Location ~? Metric Concentration txposure Selected Effect Estimates (95% Cl)
Figure 3-1 Outcome (Mg/m') Assessment'
Short-term Exposure Studies
37.
~
38.
39.
Cakmaketal.(2011)
Itoetal. (2011)
Lippmann etal.
(2006)
Ostro etal. (2007)
Zhou etal. (2011)
3 urban areas in
Chile
New York City,
NY
90 U.S. cities
and Hong Kong
6 California
counties
Seattle, WA and
Detroit, Ml
All-cause
mortality
CVD
mortality
All-cause
mortality
All-cause,
CVD and
respiratory
mortality
All-cause,
CVD and
respiratory
mortality
EC
EC
EC
EC
BC (as a
surrogate
index
of EC in
the filter
samples)
Mean (24-h avg):
2.69-5.37
Mean (24-h avg): 1.13
(Not reported)
Mean (24-h avg):
0.966
Median (24-h avg):
0.52-0.71
Continuous
measurements from
3 monitors (one in
each urban area)
Continuous
measurements from
3 monitors
Speciation data from
U.S. EPA AQS from
2000-2003
Continuous
measurements across
monitors
Continuous
measurements from
1 monitor in each city
Mortality risk ratio:
All-cause: 1.084 (1.067, 1.100)
<64yrs old: 1.052 (1.019, 1.085)
>85yrs old: 1.1 37 (1.1 03, 1.173)
Percent increase in mortality:
All year: 2.0 (0.8, 3.3)
Warm season: 2.3 (0.3, 4.3)
Cold season: 1.6 (-0.1, 3.2)
"Elevated but nonsignificant increases >0.21
[in health effect estimate] were associated
with EC"
Percent excess risk:
All-cause mortality: 0.7 (-0.6, 1.9)
Cardiovascular mortality: 2.1 (0.3, 3.9)
Respiratory mortality: 1 .2% (-2.2, 4.7)
Percent excess risk:
Detroit, warm season:
All-cause mortality: 1.359 (-1.027, 3.804)
Cardiovascular mortality: 1.227 (-2.276, 8.499)
Respiratory mortality: 0.426 (-7.046, 4.062)
Seattle, warm season:
All-cause mortality: -1.652 (-7.029, 4.035)
Cardiovascular mortality: -0.539 (-9.764, 9.629)
Respiratory mortality: 10.579 (-8.382, 33.465)
Long-term Exposure Studies
~
40.
Lipfert etal. (2006)
Lipfert etal. (2009)
Can etal. (2011)
187 U.S.
Counties
3,065 U.S.
Counties
Vancouver, BC
All-cause
mortality
All-cause
mortality
CHD
mortality
EC
EC
BC
Mean (24-h avg):
0.79
Mean (24-h avg):
0.82
Mean (24-h avg):
1.19
Speciation data from
U.S. EPA AQS from
2002
Estimates from
Atmospheric and
Environmental
Research, Inc(AER)
plume-in-grid air
quality model
High resolution land
use regression model
Coefficient (SE) for association with
mortality:
0.1664(0.05884)
Cumulative mortality risks:
All subjects: 1.07 (1.05, 1.10)
Subjects in counties with high traffic
density: 1.15 (1.13, 1.16)
Subjects in counties with low traffic density:
1.04(1.01,1.07)
Relative Risk of CHD mortality:
1.06(1.03,1.09)
co
5"
n
n
a
c-
o
3
o
3
C
c-
o'
a:
a
Er
a
3
a.
o
3
3
SI
vo
1. A more complete description or evaluation of the BC monitoring method is beyond the scope of this summary, and may include surrogate measurements.
2. SE=standard error; CHD = coronary heart disease.
-------
Chapter 3
Populations near major roads experience greater
risk of certain adverse health effects. The Health
Effects Institute (HEI) published a report on
the health effects of traffic-related air pollution
(Health Effects Institute, 2010). It concluded that
evidence is "sufficient to infer the presence of a
causal association" between traffic exposure and
exacerbation of childhood asthma symptoms. The
HEI report also concludes that the evidence is either
"sufficient" or "suggestive but not sufficient" for a
causal association between traffic exposure and new
childhood asthma cases. A review of asthma studies
by Salam et al. (2008) reaches similar conclusions.
The HEI report also concludes that there is
"suggestive" evidence for pulmonary function deficits
associated with traffic exposure, but concluded
that there is "inadequate and insufficient" evidence
for causal associations with respiratory health care
utilization, adult-onset asthma, chronic obstructive
pulmonary disease (COPD) symptoms, and allergy.
A review by Holguin (2008) notes that the effects
of traffic on asthma may be modified by nutrition
status, medication use, and genetic factors.
The HEI report also concludes that evidence is
"suggestive" of a causal association between traffic
exposure and all-cause and cardiovascular mortality.
There is also evidence of an association between
traffic-related air pollutants and cardiovascular
effects such as changes in heart rhythm, heart
attack, and cardiovascular disease. The HEI
report characterizes this evidence as "suggestive"
of a causal association, and an independent
epidemiological literature review by Adar and
Kaufman (2007) concludes that there is "consistent
evidence" linking traffic-related pollution and
adverse cardiovascular health outcomes.
Some studies have reported associations between
traffic exposure and other health effects, such as
birth outcomes (e.g., low birth weight) and childhood
cancer. The HEI report concludes that there is
currently "inadequate and insufficient" evidence
for a causal association between these effects and
traffic exposure. A review by Raaschou-Nielsen and
Reynolds (2006) concluded that evidence of an
association between childhood cancer and traffic-
related air pollutants is weak, but noted the inability
to draw firm conclusions based on limited evidence.
Investigators have attempted to trace PM health
effects back to specific sources (e.g., traffic) using
source apportionment techniques. A number
of these studies have linked BC-rich sources,
including motor vehicles and traffic, with adverse
cardiovascular and respiratory health outcomes
(U.S. EPA, 2009b, Section 6.6.2). For example,
Sarnat et a I. (2008) found consistent positive
associations between cardiorespiratory morbidity
and sources related to biomass combustion and
metal processing. However, in general there are
uncertainties associated with source apportionment
methods; these have been characterized in a recent
review (Stanek et al., 2011). First, the number of
components that comprise PM is not only large,
but the correlations between them can be high.
Some studies identify the resulting groups or factors
with named sources of ambient PM (e.g., "traffic")
or PM-related processes (e.g., "secondary organic
aerosols"), but many do not draw explicit links
between factors and actual sources or processes.
Second, there is no well-established, objective
method for conducting the various forms of
factor analysis and source apportionment, leaving
much of the model operation and assignment of
factors to sources open to judgment by individual
investigators. Because of this and differences in
composition and correlations among components
between studies, the factors identified vary
considerably, thus complicating direct comparisons.
Likewise, it cannot be ruled out that a seemingly
comparable factor across studies may correspond
to different sources depending on location. Despite
these uncertainties, a number of studies (e.g.,
Hopke et al., 2006; Thurston et al., 2005; Mar et al.,
2006; Ito et al., 2006; Sarnat et al., 2008) have found
that effect estimates based on different source
apportionment methods were generally in close
agreement, and that the variability in relative risks
across source apportionment methods was smaller
than the variability across source types (Ito et al.,
2006).
Overall, source apportionment studies report little
agreement for a particular group of components
or sources being responsible for cardiovascular
or respiratory effects, which may be due in part
to the limited number of studies evaluating these
endpoints (Stanek et al., 2011). The results of source
apportionment studies indicate that many grouped
components can be linked with various health
effects, but collectively they have not yielded a
clear and consistent association with specific health
outcomes.
Finally, it is important to note that a variety
of hazardous air pollutants (HAPs) including
polycyclic aromatic hydrocarbons (PAHs), dioxins
and furans, are co-emitted with BC (Allen et
al., 1996; Shih et al., 2008; Hedman et al., 2006;
Yadav et al., 2010; Amador-Munoz et al., 2010;
Walgraeve et al., 2010). These HAPs are associated
with adverse health effects including cancer and
respiratory effects, among others. Reductions
in HAP emissions occurring in conjunction with
BC mitigation programs will help reduce these
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Black Carbon Effects on Public Health and the Environment
health risks. Furthermore, these toxic pollutants are
generally persistent once they are emitted into the
environment, so these co-benefits can be expected
to have long-lasting beneficial impacts (Quiroz et al.,
2010; Chi et al., 2010).
3.3.2.2 Magnitude of Impacts of Ambient PM2.5 in
the United States and Globally
PM2.5 is a serious detriment to public health, both in
the United States and globally. Regulation of PM2.5
concentrations in the United States has resulted
in significant declines in PM2.5 concentrations
and PM2.5-related mortality over time (Fann and
Risley, 2011). However, many areas of the country
remain in non-attainment for the PM25 NAAQS,
and 2005 ambient PM25 concentrations have been
associated with 130,000 premature deaths annually,
corresponding to 1.1 million years of life lost (Fann
et al., 2011). While a portion of these PM2.5-related
deaths will be reduced by the recently finalized
Cross-State Air Pollution Rule aimed at controlling
SO2 and NOX emissions (U.S. EPA, 2011e), PM2.5
remains a significant risk factor for public health in
the United States.
Globally, ambient air pollution concentrations
are often much higher than those found in the
United States, and the public health burden is
correspondingly more severe. In 2004, the WHO
estimated that ambient PM2.5 in urban areas was
associated with about 800,000 premature deaths
each year globally, based on surface monitor
observations which are limited in many locations
around the world (Cohen et al., 2004). More
recently, Anenberg et al. (2010) estimated about
3.7 million global premature deaths annually
due to outdoor anthropogenic PM25 using a
global atmospheric model to isolate the total
anthropogenic contribution to PM25 concentrations
(calculated as the difference between simulated
present-day concentrations in 2000 and preindustrial
concentrations in 1860) with full spatial coverage
including both urban and rural populations. This
estimate was still considered to be an underestimate
since the resolution of the atmospheric model
was too coarse to capture fine spatial gradients of
both concentration and population, particularly
in urban areas. Impacts of outdoor PM25 were
estimated to be an order of magnitude higher than
the impacts of outdoor ozone, due both to high
PM25 concentrations, particularly in very populated
areas, and a stronger mortality relationship for PM25
relative to ozone (e.g., Jerrett et al., 2009; Krewski et
al., 2009).
The WHO estimates that urban air pollution ranks
as the 10th and 8th major risk factor in medium-
and high-income countries, respectively (World
Health Organization, 2009). Urban air pollution is
not ranked in the top 10 of major risk factors in
low-income countries since other risk factors (e.g.
childhood underweight and unsafe water, sanitation,
and hygiene) are so significant; however, a much
larger portion of deaths related to ambient PM2.5
are expected to occur in developing regions (Cohen
et al., 2004; Anenberg et al., 2010). The ongoing
Global Burden of Diseases, Injuries, and Risk Factors
Study3 is expected to update these burden estimates
leveraging the advantages from air pollution
monitors on the ground, satellite observations, and
atmospheric models.
Since the literature on differential toxicity of
PM25 components is currently inconclusive, these
studies all assume that all PM2.5 components are
equally toxic, and calculate premature deaths
associated with total PM2.5 concentrations from the
epidemiology literature. Using the same assumption,
Anenberg et al. (2011) estimated that halving
anthropogenic BC emissions globally avoids 157,000
premature deaths annually. Multiplying this estimate
by two for the total anthropogenic BC burden (using
a reasonable assumption that PM25 concentrations
respond about linearly to BC emission changes)
yields about 314,000 avoided premature deaths
annually worldwide.
3.3.3 Health Effects Related to Indoor BC
Exposures
BC is a component of indoor air pollution, which
has been implicated in an array of adverse health
effects for those who rely on solid fuels for everyday
cooking and heating, mostly in the form of biomass
(e.g., wood, animal dung, or crop wastes) but also
coal (mainly in China) (Rehfuess et al., 2006). The use
of solid fuels in poorly ventilated conditions results
in high levels of indoor air pollution, most seriously
affecting women and their youngest children (Bruce
et al., 2000; Martin et al., 2011). Recent observational
studies have suggested that indoor air pollution
from biomass fuel is associated with respiratory
morbidity, including acute lower respiratory tract
infections in children (Smith et al., 2000a; 2011) and
COPD in women (Orozco-Levi et al., 2006; Rinne et
al., 2006; Liu et al., 2007; Kiraz et al., 2003; Regalado
et al., 2006; Ramirez-Venegas et al., 2006; Ezzati et
al., 2004; Smith et al., 2004). Exposure to biomass
smoke in Guatemalan women has been shown to
increase diastolic blood pressure (McCracken et al.,
2007). Evidence also exists that implicates exposure
to biomass fuel smoke in adverse effects on
3 http://www.who.int/healthinfo/global_burden_disease/GBD_2005_
study/en/index.html.
Report to Congress on Black Carbon
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Chapter 3
different birth outcomes, including low birth weight
and stillbirth (Boy et al., 2002; Sram et al., 2005; Pope
et al., 2010). Finally, exposure to indoor air pollution
from solid fuel use has been linked to mortality
(World Health Organization, 2009).
3.3.3.1 Magnitude of Impacts of Indoor Exposures
to PM2.5 Globally
Globally, more than half of the population burns
solid fuels (e.g. coal, wood, straw, agricultural
residue, dung, etc.) for cooking and heating, mainly
in the developing world (World Health Organization,
2009). Indoor burning of solid fuels results in high
exposure concentrations, as emissions are largely
uncontrolled, the homes in which they are used
often have poor ventilation, and women and children
may spend long periods of time in direct exposure
to the emissions during cooking activities. Solid fuel
combustion emits a mixture of harmful substances,
including PM2.5. Consistent with the epidemiological
literature on indoor air pollution, impact assessments
generally relate risk of mortality with household use
of solid fuel combustion, including the total mixture
of emissions, rather than using a concentration-
response function for individual pollutants (e.g.,
PMZ5).
The WHO estimates that exposure to indoor
burning of solid fuels is associated with 2 million
annual premature deaths worldwide (World
Health Organization, 2009; Smith et al., 2004).
Globally, indoor smoke from solid fuels ranks as
the 10th leading risk factor for premature death
and contributes 3.3% of total deaths. In terms of
overall disease burden, as measured in Disability
Adjusted Life Years (DALYs), indoor smoke from solid
fuels ranks as the 9th leading risk factor globally,
associated with 2.7% of all DALYs. It is particularly
a problem in low-income countries, where indoor
smoke from solid fuels ranks as the 6th leading
mortality risk factor (4.8% of total deaths) and
the 5th leading disease risk factor (4% of all DALYs).
Indoor smoke from solid fuels does not rank as a
major risk factor for high-income countries, where
use is relatively limited and ventilation is generally
sufficient to maintain air quality indoors. As for
ambient air pollution, the ongoing Global Burden of
Diseases, Injuries, and Risk Factors Study4 is expected
to update these burden estimates with improved
assumptions and more recent demographic
information.
3.4 Non-Climate Welfare Effects of
PM2 5, Including BC
Non-climate welfare effects resulting from BC
emissions are discussed in terms of PM2.5 exposure
and deposition. Visibility impairment, which is
caused by light scattering and absorption by
suspended particles and gases, is the primary non-
climate welfare effect of BC. Crop yields may also be
adversely affected by exposure to and deposition
from PM2.5. PM2.5 has been linked to adverse impacts
on ecosystems, primarily through deposition of
PM constituents. In addition, deposition of PM is
associated with damages to materials and buildings.
3.4.1 Role of BC in Visibility Impairment
Particles are the dominant air pollutant responsible
for visibility impairment, e.g. "haze," in both urban
and remote areas. In the same way that particles
influence the Earth's radiative balance, by scattering
and/or absorbing solar radiation, they influence
the quantity and quality of light received by the
human eye and, therefore, one's ability to recognize
and appreciate the form, contrast detail, and color
of near and distant features. Aerosol-based light
extinction can be estimated using the Interagency
Monitoring of Protected Visual Environments
(IMPROVE) algorithm that multiplies the ambient
concentration of PM components by typical
component-specific light extinction efficiencies.5
BC and crustal minerals are the only included
components that contribute to light absorption.
Under low humidity conditions, BC and OC have the
greatest effect on visibility among the major PM
species. Per unit mass, the algorithm specifies that
BC is 2.5 times more effective at absorbing light
than organic carbon is at scattering.
Carbonaceous PM is responsible for a large fraction
of regional haze, particularly in the Northwest,
where annual average concentrations for 2000-2004
account for 40-60% of the aerosol based light
extinction. Most of this average carbonaceous
visibility impairment throughout the United States
is associated with OC (in both rural and urban
5 See http://vista.dra.colostate.edu/improve. For two major PM2.s
components, sulfate and nitrate, water growth factors are included
to account for enhanced light extinction due to relative humidity.
The original IMPROVE equation included Rayleigh scattering (from
natural atmospheric gasses) and factors for particulate sulfates,
nitrates, organic carbon, elemental carbon, fine soil and coarse
particles, with a hygroscopic growth function for enhanced light
scattering from water associates with the sulfates and nitrates. A
recently proposed revision to this equation (Pitchford et al., 2007)
enhances the scattering from high concentrations of sulfates,
nitrates or organics and adds terms for scattering and hygroscopic
growth from sea salt and for light absorption from gaseous NO2.
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Black Carbon Effects on Public Health and the Environment
areas) because of relatively high OC concentrations
compared to BC. Regional haze in the eastern United
States generally contains even higher concentrations
of carbonaceous PM and light-absorbing BC plays a
relatively larger but still minor role compared to OC
(DeBell, 2006).
As described further in Chapter 5, urban areas have
more carbonaceous PM than nearby remote (rural)
areas in the same region (U.S. EPA, 2004b). Western
urban areas have more than twice the average
concentrations of carbonaceous PM than remote
areas in the same region (DeBell, 2006). As shown
in Figure 5-6, average urban PM2.5 is composed
of roughly equal proportions of carbonaceous
and sulfate components in some eastern areas. In
conditions of high relative humidity common in the
eastern United States, hydrated sulfate dominates as
the constituent responsible for most urban haze on
the haziest summer-time days (U.S. EPA, 2009b).
The 1977 Clean Air Act Amendments called for the
development of regulations to address regional
haze (visibility impairment) in 156 National Parks
and wilderness areas in the United States. The EPA
promulgated a Regional Haze Rule (RHR) in 1999 in
response to this mandate. Implementation of the
RHR entails planned emissions reductions to ensure
that by 2064, the worst haze days in these protected
areas will improve to natural conditions without
degrading visibility conditions for the best haze days.
In addition to the RHR aimed at achieving visibility
improvements in protected National Park areas, the
NAAQS program has been successful at achieving
visibility improvements in rural areas, as well as in
urban areas where people live and work.
3.4.2 Role of BC in Crop Damage and Other
Environmental Impacts
Crop yields can be sensitive to the amount of
sunlight received. As discussed in detail in Chapter 2,
BC and other airborne particles contribute to surface
dimming, and crop losses have been attributed
to increased airborne particle concentrations in
some areas of the world (Chameides et al., 1999).
Auffhammer et al. (2006) found that fossil fuel
and biomass burning contributes to reduced rice
harvests in India. Decreases in rice and winter wheat
yields have also been attributed to regional scale air
pollution in China (Chameides et al., 1999).
Ecological effects of PM include direct effects to
metabolic processes of plant foliage (Naidoo and
Chirkoot, 2004; Kuki et al., 2008); contribution to
total metal loading resulting in alteration of soil
biogeochemistry (Burt et al., 2003; Ramos et al.,
1994; Watmough et al., 2004), plant growth (Audet
and Charest, 2007; Kucera et al., 2008; Strydom et al.,
2006) and animal growth and reproduction (Gomot-
de Vaufleury and Kerhoas, 2000; Regoli et al., 2006);
and contribution to total organics loading resulting
in bioaccumulation and biomagnification across
trophic levels (Notten et al., 2005).
Building materials (metals, stones, cements, and
paints) undergo natural weathering processes
from exposure to environmental elements (wind,
moisture, temperature fluctuations, sunlight, etc.).
Deposition of PM is associated with both physical
damage (materials damage effects) and impaired
aesthetic qualities (soiling effects) for building
materials. Wet and dry deposition of PM can
physically affect materials, adding to the effects
of natural weathering processes, by potentially
promoting or accelerating corrosion of metals, by
degrading paints and by deteriorating building
materials (Haynie, 1986; Nazaroff and Cass, 1991).
Fine particles may coat building materials, damaging
the appearance of homes, public buildings, and
historic landmarks (Hamilton and Mansfield, 1991).
Studies have been conducted by a number of
authors identifying the anthropogenic sources of
soiling and materials damages to monuments and
historical buildings (Sabbioni et al., 2003; Bonazza
et al., 2005). For example, Bonazza evaluated
deposition to the London Tower and found that
"deposition of elemental carbon darkens surfaces
and has importantly aesthetic implications for
buildings." Reduction of PM deposition is beneficial
in terms of reduced cleaning, maintenance,
and restoration expenditures for buildings and
structures.
3.5 Key Uncertainties Regarding
Health/Environmental Impacts of BC
A review of the literature describing the health
effects associated with ambient concentrations
of BC indicates that the strongest relationship
exists between BC and cardiovascular effects.
This evidence includes support for a continuum
of cardiovascular effects ranging from subtle
subclinical measures to more severe effects on
the cardiovascular system, such as emergency
department visits and hospital admissions. These
associations are generally consistent with the
associations observed for PM2.5 and cardiovascular
effects (Janssen et al., 2011), though the body of
evidence describing the association between BC
and cardiovascular effects is much smaller and
less consistent than the one characterizing PM2.5
and cardiovascular effects. It is noteworthy that,
among the studies that characterize the association
between BC and cardiovascular effects, a large
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Chapter 3
majority have been conducted in the greater Boston,
MA area, utilizing BC measurements from a single
BC monitor (see Table 3-2). There can be substantial
spatial variation in BC concentrations within a single
city, and ambient concentrations of BC in any urban
area can vary widely from location to location within
the city. Thus, the reliance on this single monitor to
estimate exposure for a number of studies across
the entire greater Boston, MA area may contribute
uncertainty to the reported associations. Similarly,
the ambient concentration and composition of PM is
geographically heterogeneous, with variations due
to unique PM sources and from unique formation,
transport, transformation, removal, and infiltration
processes in different locations. Thus, a body of
evidence that is focused on one geographic area,
in this case Boston, MA, introduces uncertainty to
the characterization of the association between
ambient BC and cardiovascular effects, and the
generalizability of this association to broader
geographic areas.
An additional uncertainty regarding the health
impacts of BC is the inconsistency between the
results of studies examining ambient concentrations
of BC and the results attributed to traffic in the HEI
report (Health Effects Institute, 2010). In examining
the body of evidence for health effects associated
with BC, the strongest relationship was observed for
BC and cardiovascular effects, while the evidence
for an effect of BC on respiratory effects was
observed to be inconsistent. Conversely, the HEI
report on traffic (Health Effects Institute, 2010),
concluded that evidence is "sufficient to infer the
presence of a causal association" between traffic
exposure and respiratory effects (i.e., exacerbation
of childhood asthma symptoms), while the evidence
for an association with cardiovascular effects was
"suggestive." Thus, while BC is a known component
of the air pollution mixture attributed to traffic
sources, it may have a stronger association with
some health effects attributed to traffic (i.e.,
respiratory effects) than others (i.e., cardiovascular
effects). Furthermore, this line of reasoning indicates
that there are likely additional components to the
air pollution mixture attributed to traffic sources
(other than BC) that contribute to the health effects
associated with exposure to traffic. Additionally,
BC could be serving as an indicator for a larger
category of primary combustion particles, which,
in addition to BC, can include trace metals and
hydrocarbons such as PAHS, any or all of which
could be acting to cause adverse health effects.
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Chapter 4
Emissions of Black Carbon
4.1 Summary of Key Messages
• Emissions of BC from U.S. sources total about 0.64
million tons (580 Gg) in 2005, which represents
about 8% of the global total. Mobile sources
account for a little more than half (52%) of the
domestic BC emissions. Approximately 93% of
the mobile source total is from diesel sources.
Open biomass burning is the next largest source
in the United States, accounting for about 35% of
the total. In general, BC is concentrated in urban
areas, where populations are largest, making
health an important issue in addition to climate in
BC mitigation strategies.
• OC is a significant co-emitted pollutant among
the major BC emitting sources. The United States
is estimated to emit about 1.7 million tons (1500
Gg) of OC. The ratio and mass of BC and OC varies
by source. Diesel combustion emissions produce
the largest fraction of BC while emissions from
open biomass burning are dominated by OC.
More research is needed on how OC/BC ratios can
be used to characterize the net climate impacts of
different sources.
• Diesel sources have a low OC/BC ratio, making
them strong candidates for mitigation. By 2030,
domestic diesel emissions will be reduced by
the phase-in of recent national mobile source
emission standards, and other categories, such
as open biomass burning, will emerge as top
emitters of BC in the United States.
• More than two-thirds of the almost 8.4 million
tons (7,600 Gg) of global BC emissions come from
open biomass burning and residential sources.
The regions of the world responsible for the
majority (nearly 75%) of BC emissions world-wide
are Africa, Asia, and Latin America. In developing
countries, biomass burning and residential
sources are the dominant sources of BC, while in
developed countries, emissions of BC are lower
and are often dominated by transportation and
industry.
• Long-term historic trends of BC emissions in
the United States reveal a dramatic increase
in emissions from contained combustion
sources from the mid-1880s to approximately
the 1920s followed by a decline over the next
eight decades. The decline can be attributed to
changes in fuel use, more efficient combustion
of coal, and implementation of PM controls.
In contrast, developing countries (e.g., China
and India) have shown a very sharp rise in BC
emissions over the past 50 years.
• Characterization of domestic and global BC
emissions and the subsequent development of
BC emissions inventories are based on a limited
number of existing source measurements. Better
information is needed on chemical composition
of PM for some critical emissions sources to
improve estimates of BC in these inventories.
4.2 Introduction
Emissions inventories provide valuable information
about major sources of BC, both domestically and
internationally, and the trends in BC emissions
over time. This chapter covers domestic and global
emissions of BC and OC. In the case of domestic
emissions, the discussion begins with source
measurements that generate speciated emissions
profiles and ends with a description of the current
U.S. emissions inventory for BC and OC by source
category, with particular attention to mobile
sources, open biomass burning, and fossil fuel
combustion.1 This chapter also provides an overview
of key emissions estimates from available global
inventories as well as inventories for key world
regions such as China and India, and evaluates
historical trends in global emissions. This chapter
includes a comparison of the U.S. portion of the
global BC inventory to the EPA developed estimates.
In addition, this chapter discusses the implications of
long-range transport of aerosols, which contributes
to total BC in the column of air above an area. Based
1 Most estimates of source emissions in the United States
utilize thermal optical methods which estimate BC as elemental
carbon (EC). However, for purposes of this chapter, all emissions
estimates will be referred to as BC. This issue is addressed for
ambient measurements in Chapter 5 and covered in more detail in
Appendix 1.
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85
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Chapter 4
on the discussion in this chapter, key emissions
research needs for BC and other light absorbing
aerosols are incorporated into the recommendations
discussed in Chapter 12.
4.3 U.S. Black Carbon Emissions
4.3.1 Summary of Emissions Methodology
Currently, the U.S. EPA does not require the states to
report emissions of BC and other PM constituents
(OC, nitrates, sulfates, crustal material) as part of
the National Emissions Inventory (NEI). Rather,
the U.S. emissions inventory uses total PM2.5
emissions to derive estimates for direct emissions of
carbonaceous particles, including BC and OC, for all
sources except on-road mobile sources. Therefore,
all of the available emissions inventory information
on carbon emissions in the United States is restricted
to those source categories with sufficient PM2.5
emissions estimates to support this derivation. The
Natural Gas Combustion
Process Gas Combustion -
Wood Fired Boiler- |
Distillate Oil Combustion
Agricultural Burning- -[]
Wildfires- •$•
Residential Wood Combustion-
SubBituminous Combustion f
Charbroiling- fl-
Prescribed Burning 0-
Bituminous Combustion- Q-
0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 0.0 0.2 0.4 0.6 0.8
Figure 4-1. BCand OC Fractions of PM25 Emissions for the Highest BC Emitting
Non-Mobile Source Categories in the United States. The box represents the 25th
to 75th percentile range and the whiskers represent the 10th and 90th percentile
points of the individual source profiles based on emissions source test data as it
exists in EPA's SPECIATE database for each source category. The vertical lines within
the box represent the median values for that source category. The dots represent
outliers. Some of the outliers show a fraction greater than unity; that is due to the
statistical procedures used to composite averages. See Appendix 2 for further
details. (Source: U.S. EPA)
methods used to generate U.S. emissions inventories
are described in detail in Appendix 2.
In general, EPA estimates emissions of BC and OC by
appropriately matching PM2.5 emissions estimates
from EPA's NEI with source profiles contained
in EPA's SPECIATE database (see Appendix 2 for
details). SPECIATE is the EPA's repository of PM and
VOC speciation profiles of air pollution sources.
The PM speciation profiles contain weight fractions
of chemical species (e.g., BC and OC) for specific
sources. Applying these profiles to PM emissions
inventories provides estimates of how much BC
and OC is emitted by specific source categories.
There are about 300 profiles in the SPECIATE
database that are of sufficient quality for this
purpose. The mapping of how these approximately
300 profiles have been applied to the over 3,400
source categories available in EPA's NEI for PM2.5
is described in Appendix 2 and more details are
available in the literature (Reff et al., 2009; Simon
et al., 2010). For all non-mobile source and non-
open biomass emissions
estimates, all BC and
OC estimates are based
oc on EPA's 2005 modeling
inventories (termed
1 "2005CK" inventories),
which rely on the 2005
NEIforPM2.5.
As noted above, for
on-road mobile source
categories (e.g., cars and
trucks), BC is predicted
directly without using
SPECIATE. For on-road
gasoline and diesel
vehicles, emissions
estimates are generated
directly through models.
Appendix 2 provides
details on how these
emissions were calculated
using EPA emissions
models for on-road and
nonroad vehicles/engines,
and also discusses other
important issues, like high
emitters, deterioration
of PM emissions (i.e.,
increase in PM mass)
with age, and increased
PM emissions at lower
temperatures. All three of
these issues are important
and available data on
them are incorporated
-m-
-m-
-D-
86
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Emissions of Black Carbon
Metals & Elements
2%
Sulfate, Nitrate (1-5%)
1%
(1-4%)
Other
3%
(1-10%)
Organic Carbon
19%
(7-49%)
Elemental Carbon
75%
(33-90%)
Figure 4-2. Heavy-Duty Diesel PM2.5 Emissions Profile.
(Source: U.S. EPA, 2002b)
into EPA's emissions models. There are more data on
these issues for gasoline PM than for diesel PM.
PM2.5 emissions from open biomass burning
(wildfires, agricultural burning, and prescribed
burning) come from an emissions inventory
compiled by the Regional Planning Organizations
(RPOs) for the year 2002 (Regional Planning
Organization, 2004a; 2004b; 2005; 2006; 2008).
There are five RPOs in the United States which are
set up to address regional haze and related issues
across the country. Due to the need to accurately
represent local/regional fire emissions, each RPO
has invested time in including greater regional/
local specificity resulting in development of more
accurate fire inventories, thereby making them more
accurate than national estimates developed by EPA.
In addition, these RPO estimates have received more
widespread review and acceptance by the states,
RPOs and other federal agencies. Though these
emissions estimates represent the year 2002, the
difference between the year of estimates matters
less than the accuracy and review of the estimates;
this is because there is very little year-to-year
variation in categories besides wildfires. In the case
of wildfires, these 2002 estimates are consistent with
an average of wildfire activity over a ten year period
from 2001 to 2010. BC and OC emissions were then
estimated based on these PM2.5 estimates using
the same methodology explained above. Despite
the higher accuracy of RPO emissions estimates as
compared to EPA's, it should be noted that biomass
burning BC estimates remain more uncertain than
engine combustion BC, for example, because of the
tremendous year-to-year variability in open burning
activity and for other reasons addressed later in this
chapter.
It is also important to note that
the BC and OC inventories do not
account for secondary formation
of particles in the atmosphere.
While secondary formation is not
substantial for BC, a significant
amount of OC can be formed in
the atmosphere from biogenic
and anthropogenic emissions of
volatile organic chemicals. Most air
quality and climate models rely on
estimates of OM (which is OC plus
the mass that accrues to primary
OC through photochemistry in the
atmosphere), rather than OC, to
calculate atmospheric reactions and
impacts.
Figure 4-1 displays the number of
resulting profiles (the numbers on
the right-hand side of the graphs)
and their distribution of BC and OC fractions of
PM2.5 by source category. Mobile source categories
are not shown in Figure 4-1 due to the fact that BC
emissions are estimated directly from models for
some of those sources (see below and Appendix 2
for more information on all mobile source
categories). The number of individual profiles by
source category can be quite limited; sometimes
only a single value is known. The source categories
depicted on the y-axis in Figure 4-1 are shown
top-to-bottom in order of maximum BC fraction
to minimum BC fraction. Natural-gas combustion
(see caveat later in this chapter) has only one profile
available but has the highest BC fraction of the
source categories shown in Figure 4-1 while some
of the burning and coal combustion categories have
the lowest BC fractions.
Heavy-duty diesels have the highest BC fraction
of all source categories, at an average of about
77% although this percentage varies depending
on operating mode and engine technologies. This
fact is supported by the EPA's Health Assessment
Document for Diesel Engine Exhaust (2002b) in
which the chemical composition of diesel engine
exhaust is identified as shown in Figure 4-2, with
BC contributing 75% of the total PM2.5 composition.
Light-duty gasoline vehicles have a much smaller
fraction (about 20-25%) of PM that is BC.
4.3.2 U.S. Black Carbon Emissions:
Overview and by Source Category
In 2005, the United States is estimated to have
emitted about 5.5 million tons (or about 5,000
Gg) of primary PM2.5 of which about 0.64 million
tons (12%) was BC and about 1.7 million tons
Report to Congress on Black Carbon
87
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Chapter 4
U.S. PM2.5 Emissions in 2005 (5.5 Million Tons)
Open Biomass Burning
(Includes Wildfires)
41.0%
Residential/Domestic
8.4%
U.S. BC Emissions in 2005 (0.64 Million Tons)
Transport/Mobile
52.3%
Open Biomass Burning
{Includes Wildfires}
35.3%
Residential/Domestic
3.6%
(30%) was primary OC.2'3'4 Thus at a national level,
more than twice as much OC is emitted from
domestic sources as BC. The domestic emissions
of 0.64 million tons represents about 7% of the
world's total BC emissions (i.e., 8.4 million tons)
making the United States the 8th largest global
BC emitter (Lamarque et al., 2010). The majority
of U.S. BC emissions come from mobile sources
(predominantly diesel) and open biomass burning.
In 2005, about 65% of total U.S. BC was emitted in
urban counties and, in the case of mobile sources,
more than 70% of the total U.S. BC emissions occur
in urban counties. In addition, it should be noted
that all emissions estimates shown in this chapter
are annual averages. There is expected to be some
seasonal patterns in BC emissions for some of the
source categories. However, that detail is beyond
the scope of this chapter. All emissions numbers are
only estimates, as detailed in the Appendices. There
are uncertainties in all of these estimates that vary
from category to category. Systematic quantitative
estimates of uncertainty in U.S. emissions estimates
are not available at this time, though some
qualitative discussion is provided both in this
chapter and in the Appendices.
Figure 4-3 displays the percentage of total U.S.
emissions of primary PM2.5, BC, and OC for six
"mega" source categories:
• Open biomass burning (agricultural burning,
wildfires, and prescribed burning)
• Residential (any combustion for residential
activities regardless of fuel burned)
U.S. OC Emissions in 2005 (1.7 Million Tons)
Transport/Mobile
12.3%
Industry
1.0%
Residential/Domestic
12.3%
Open Biomass Burning
(Includes Wildfires)
63.7%
Figure 4-3. Contribution to Primary PM2.5, BC, and OC
Emissions by Mega Source Categories. (Source: U.S. EPA,
2002a, 2005a)
Energy/power (EGUs and other power generation
sources)
Industrial
2 The U.S. emissions estimates presented in this chapter reflect data
from EPA's National Emissions Inventory and mobile source models,
supplemented with data from U.S. Regional Planning Organizations
(RPOs) on open biomass burning (wildfires, agricultural burning,
and prescribed burning). Most estimates presented in this chapter
are for the year 2005. However, all emissions estimates for open
biomass burning are based on a 2002 inventory developed by the
RPOs, which are partially funded by EPA. For ease of reference,
these various sources are grouped under the label "U.S. EPA" in
figures and tables throughout this chapter. More detail on how
different portions of the inventory are constructed is provided in
other parts of this chapter and in Appendix 2.
3 Unless otherwise specified, the term "tons" refers to short tons
throughout this report. 1102 short tons = 1 Gigagram.
4 This does not account for other components of organic PM
emissions, such as oxygen and hydrogen.
Report to Congress on Black Carbon
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Emissions of Black Carbon
Table 4-1. 2005 U.S. Emissions (tons) and Ratios of Emissions by Mega Source Category. (Source: U.S. EPA)
Mega Source Category 1 PM2.5 1 BC 1 OC 1 OC/BC 1 BC/PM2.5
Open Biomass Burning
Residential
Energy/Power
Industrial
Mobile Sources (Transport)
Other
Totals (Short Tons)
In GigaGrams (Gg)
2,266,513
464,063
712,438
219,460
626,859
1,232,123
5,521,456
(5,009)
224,608
22,807
43,524
6,085
333,400
6,743
637,167
(578)
1,058,494
204,160
65,138
16,234
205,172
112,967
1,662,165
(1,508)
4.7
9
1.5
2.7
0.6
16.8
2.61
0.1
0.05
0.06
0.03
0.53
0.01
0.12
• Transport/mobile (including on-road, nonroad,
locomotives, commercial marine, aircraft and tire/
brake wear)
• Other
Table 4-1 shows the actual tons per year of direct
PM2.5, BC, and OC emissions for these source
categories, as well as some key emissions ratios. In
the last row, emissions in Gigagrams (Gg) are shown
in parenthesis, since metric units are standard for
reporting global emissions.5
Figure 4-3 clearly shows mobile sources are the
dominant contributor to total BC emissions in the
United States in 2005. Mobile sources contribute 52%
of the total BC emissions, followed by open biomass
burning (35%),6 and energy/power (7%). All other
categories are about 4% or less. Additional detail
on the specific sources that comprise these mega
source categories is provided later in this section.
As shown by the ratios in Table 4-1 (OC/BC and BC/
PM2.5), the composition of primary PM2.5 emissions
varies significantly among source categories. As
discussed in Chapter 2, such differences have
important implications for climate. For example,
5 In global inventories, total emissions are often grouped into two
main categories, "open" vs. "contained" (or "closed"). To avoid
confusion among these terms in this report, the term "contained
combustion" is used to refer to all sources except open biomass
burning. This is consistent with the global emissions inventory
literature. "Contained combustion" is a broadly encompassing
term, referring to all combustion sources in which fuel is burned
in a chamber or controlled environment (including sources such
as industrial/EGU boilers, internal combustion engines, stationary
diesel engines, and contained burning of biomass in sources such as
wood-fired boilers).
6 This total includes wildfires. The distinct contributions of wildfires
and agricultural/prescribed fires to total domestic emissions of
primary PM2.5, BC and OC are provided in Table 4-2.
diesel-powered mobile sources emit significantly
more BC than OC, while the opposite is true for
open biomass burning and residential sources.
Figure 4-4 displays the total BC emissions for the
different source categories. The data in Table 4-1
also show that for some source categories, BC and
OC together make up less than 50% of total PM2.5
emissions, indicating that there are significant
amounts of other/unidentified primary co-pollutants
(such as direct emissions of nitrates and sulfates) in
the emissions mixture.
The mega source categories can be subdivided
into more specific categories. Table 4-2 shows the
national-level emissions of primary PM2.5, BC, and
OC emissions for about 90 specific sub-categories
Figure 4-4. U.S. BC Emissions (tons) for Major Source
Categories. (Source: U.S. EPA)
Report to Congress on Black Carbon
89
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Chapter 4
Table 4-2. U.S. Emissions of PM2.5, BC, and OC (short tons). (Source: U.S. EPA)
M <; Total
Gate o"rCe Specific Category Primary BC Primary OC OC/BC BC/PM2.5
PM2.5
Open Biomass
Residential
Energy/Power
Industrial
Wildfires
Prescribed Burning
Agricultural Burning
Residential Wood Combustion
Residual Oil Combustion
Residential Coal Combustion
Residential Natural Gas Combustion
Natural Gas Combustion
Bituminous Combustion
Sub-Bituminous Combustion
Distillate Oil Combustion
Wood Fired Boiler
Process Gas Combustion
PMS02 Controlled Lignite
Combustion
Stationary Diesel
Cement Production
Ind Manuf- Avg.
Mineral Products - Avg
Kraft Recovery Furnace
Chem Manuf- Avg
Lime Kiln
Heat Treating
Aluminum Production
Ferromanganese Furnace
Surface Coating
Cast Iron Cupola
Electric Arc Furnace
Secondary Aluminum
Sintering Furnace
Pulp & Paper -Avg
Catalytic Cracking
Secondary Copper
Ammonium Nitrate Production
Secondary Lead
Petroleum Ind -Avg
Copper Production
Ammonium Sulfate Production
Open Hearth Furnace
1,600,358
535,627
130,528
379,878
78,672
2,648
2,865
64,239
394,853
143,383
23,718
56,289
9,457
20,499
4,476
17,523
46,501
23,632
21,222
17,526
7,002
14,439
5,730
1,240
9,165
3,479
4,317
6,057
5,739
6,569
8,864
1,137
1,025
410
6,224
432
65
6,686
151,855
58,525
14,228
21,194
787
634
192
24,668
6,697
6,028
2,372
2,088
1,378
293
3,452
514
416
347
325
320
162
144
132
125
64
32
16
12
10
7
6
1
738,997
268,826
50,671
200,645
787
1,187
1,541
15,867
10,387
4,514
5,930
19,764
2,850
5,826
786
2,221
3,422
1,242
1,111
1,608
466
1,011
223
64
1,903
222
140
91
157
1
11
218
1,337
4.9
4.6
3.6
9.5
1
1.9
8
0.64
1.6
0.7
2.5
9.5
2
19.9
0.2
4.3
8.2
3.6
3.4
5
2.9
7
1.7
0.5
29.7
6.9
8.8
7.6
15.7
0.2
11
0.09
0.11
0.11
0.06
0.01
0.24
0.07
0.38
0.02
0.04
0.1
0.04
0.15
0.01
0.77
0.03
0.01
0.01
0.02
0.02
0.02
0.01
0.02
0.1
0.01
0.01
0
0
0
0
0
0
0
0
0
0
0
0
90
Report to Congress on Black Carbon
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Emissions of Black Carbon
Mega Source Specific Category Primary BC Primary OC OC/BC BC/PM2.5
a egory PM2.5
Mobile Sources
Other
On-road diesel
Nonroad diesel
Locomotive
Commercial Marine (C1 & C2)
On-road gasoline
Nonroad gasoline
Commercial Marine (C3)
Tire
Brake wear
Aircraft
Charbroiling
Wood Products - Drying
Paved Road Dust
Dairy Soil
Wood Products-Sawing
Overall Average Manufacturing
Unpaved Road Dust
Charcoal Manufacturing
Solid Waste Combustion
Wood Products - Sanding
Asphalt Manufacturing
Fiberglass Manufacturing
Agricultural Soil
Fly Ash
Phosphate Manufacturing
Industrial Soil
Food &Ag - Handling
Urea Fertilizer
Potato Deep-Frying
Glass Furnace
Calcium Carbide Furnace
Sludge Combustion
Crustal Material
Brick Grinding and Screening
Auto Body Shredding
Inorganic Fertilizer
Asphalt Roofing
Limestone Dust
Sand & Gravel
Construction Dust
Meat Frying
Lead Production
208,473
145,289
30,910
28,119
75,924
55,834
56,028
5,325
17,801
3,156
64,124
8,113
54,481
9,862
12,355
10,577
419,648
5,578
14,965
2,257
2,160
4,641
334,515
1,733
992
2,011
10,331
589
192
7,803
314
163
1,160
1,272
129
78
1,872
1,912
134,885
96,669
12,216
33
153,477
112,058
22,495
21,652
14,510
5,444
1,681
1,198
475
410
2,601
649
569
509
469
466
409
290
228
135
124
93
67
30
27
23
18
12
8
5
4
2
2
1
1
1
0
44,423
30,618
5,130
4,937
59,657
46,734
6,303
3,060
2,321
1,988
42,975
4,057
5,308
3,139
5,498
927
22,897
100
1,258
790
93
1,299
10,310
21
78
20
418
183
121
55
23
14
62
31
10
2
1,129
4,463
7,012
0.3
0.3
0.2
0.2
4.1
8.6
3.7
2.6
4.9
4.8
16.5
6.3
9.3
6.2
11.7
2
56
0.3
5.5
5.9
0.8
14
153.9
0.7
2.9
0.9
23.2
15.3
15.1
11
5.8
7
31
31
10
2
0.74
0.77
0.73
0.77
0.19
0.1
0.03
0.22
0.03
0.13
0.04
0.08
0.01
0.05
0.04
0.04
0
0.05
0.02
0.06
0.06
0.02
0
0.02
0.03
0.01
0
0.02
0.04
0
0.01
0.02
0
0
0.01
0.01
0
0
0
0
0
0
Report to Congress on Black Carbon
91
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Chapter 4
M^eS°"rce Specific Category Primary BC Primary OC OC/BC BC/PM2.5
PM2.5
Other
Synthetic Residential Wood
Combustion
Sandblasting
Steel Desulfurization
Inorganic Chemical Manufacturing
Gypsum Manufacturing
Food &Ag-Drying
Boric Acid Manufacturing
Coke Calcining
Sea Salt
345
1,673
259
4,161
1,395
5,551
11
811
287
8
666
0
0
0
0
0
0
0
0
0
Notes:
1. All emissions are for 2005 except those for "open biomass burning," which are based on 2002 RPO estimates (as referenced earlier).
2. This table represents all emissions in column D as BC; however, they were derived from thermal-optical monitoring techniques and
reported as EC.
3. Blank cells indicate that the profiles used showed no BC emissions from these sources.
4. Aircraft inventories only include emissions from landings and take-offs and do not include in-flight emissions.
5. In this table, the mobile source inventories are for all 50 states. Wildfire emissions are for the 48 contiguous states plus Alaska. All
other estimates are only for the 48 contiguous states (AK and HI are expected to be m inor BC and OC contributors for all these
sources).
6. BC emissions from "Agricultural Burning" are very dependent on the types of burning activity included (e.g., range land, crop residue,
and other types of burning activity). The data used in this report to characterize prescribed burning includes range land and crop
residue burning activity as well as other types. Other recent work using satellite-imagery shows the total "agricultural emissions" in
the United States (averaged over 5 years) are somewhat lower than the BC emissions estimates shown here (McCarty, 2011). McCarty's
estimates for BC emissions from agriculture burning are based on the inclusion of crop residue burning only and PM emission factors
for that type of burning. This is a limited definition of "agricultural burning" that others also feel is appropriate. Working with USDA,
EPA is in the process of evaluating this work as well as more of its own recent work on a 2008 fires inventory that relies on updated
remote sensing methods to estimate emissions from agricultural burning.
of sources in the United States. Table 4-2 also
shows OC/BC and BC/PM2.5 ratios for each of the
specific source categories. Some of these data are
drawn from the NEI, EPA's "bottom-up" compilation
of estimates of air pollutants discharged on an
annual basis and their sources (U.S. EPA, 2005a). As
discussed previously, the "open biomass burning"
categories shown in yellow come from an emissions
inventory compiled by the RPOs for the calendar
year 2002 (Regional Planning Organization, 2004a;
2004b; 2005; 2006; 2008).
4.3.2.1 Emissions from Mobile Sources
Mobile sources account for about 52% of total U.S.
BC emissions in 2005. Within this category, emissions
from diesels (both nonroad and on-road) dominate,
accounting for about 93% of BC. Gasoline vehicles/
engines are responsible for most of the remaining
BC emissions from the mobile source category.
Figure 4-5 shows this more detailed breakout of
mobile source BC emissions. In general, diesel PM2.5
Nonroad gasoline
16% Commercial marine (C3)
0.5%
On-road gasoline
4.4%
Commercial marine
(C1 & C2)
6.5%
Locomotive
6.7%
~T're Brakewear
0.4% Q.1%
Nonroad diesel
33.6%
On-road diesel
46.0%
Figure 4-5. U.S. BC Emissions from all Mobile Source
Categories (333,400 tons). (Source: U.S. EPA)
92
Report to Congress on Black Carbon
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Emissions of Black Carbon
consists of about 70-80% BC and about 20% OC.7
Gasoline PM2.5, in contrast, consists of about 20%
BC with the remainder being mostly OC. The BC/
PM fraction though varies by operating condition
such as cold start, high speed/load, and normal
driving. Also, as discussed in Appendix 2, diesels
used for large (ocean-going) commercial marine
vessels have much lower BC/PM fractions due to
the high amount of organics and sulfates. Diesel PM
is unique in having a very high ratio of BC to OC.
The total light absorbing capacity of the specific
compounds and the resultant mixture emitted in
diesel or gasoline exhaust is not known. However,
several mobile source measurements suggest that
particle emissions from both gasoline and diesel
vehicles are strongly light absorbing (Japar et al.,
1984; Strawa et al., 2010; Adler et al., 2010; Japar and
Szkarlat, 1980). It should be noted that while mobile
sources represent about 52% of the national total of
BC emissions, they represent about 69% of all non-
wildfire BC emissions in the United States.
While mobile sources dominate the U.S. inventory
currently, significant reductions in emissions of
both BC and OC have been achieved since 1990,
and existing vehicle regulations are expected to
produce further reductions in coming years as they
are implemented. Most of these BC reductions are
a direct result of EPA's regulations on diesel PM, but
reductions in total carbon emissions, mostly OC, are
also due to regulations on emissions from gasoline
vehicles. Due to these regulations, the mobile source
contribution to BC compared to other sources has
declined on both an absolute basis and a fractional
basis since 1990. As reductions continue through
2030 and beyond, the pie chart shown in Figure 4-3
will continue to change, showing an increasingly
smaller contribution of mobile sources to overall U.S.
BC emissions. Chapter 8 summarizes mobile source
BC inventories for various years from 1990 through
2030, and the control programs that are expected to
result in these emissions reductions by 2030.
Numerous source apportionment studies have been
done for local areas in the United States including
Denver, Los Angeles, Atlanta, Phoenix, and other
areas (Watson et al., 1998; Zheng et al., 2007; Brown
et al., 2007). These studies show the importance
of mobile source emissions to ambient PM2.5 (and
in some cases, BC) levels, just as various emissions
inventories do. The inventories in this report are
nationwide inventories and, thus, are not comparable
to source apportionment studies which are done
in local areas. Also, there are various workshops
and studies looking at differing health impacts of
PM components including some health studies
on source apportionment (Thurston et al., 2005;
Marmuretal., 2006).
4.3.2.2 Emissions from Biomass Combustion
Several source categories in Table 4-2 include
emissions from "wood based" (biomass) combustion.
Based on an approach suggested by Bond et
al. (Bond, 2007; Bond et al., 2004) to facilitate
consideration of mitigation options, elements of
these source categories: "open biomass burning",
"residential heating/cooking", and "biomass fired
stationary sources" have been combined into a
"biomass combustion" category for this discussion.
Table 4-3 summarizes the sources included this
"biomass combustion" category and their associated
emissions.
These biomass combustion sources are estimated to
collectively emit a little more than 250,000 tons of
BC annually. This represents about 39% of the total
amount of BC emitted in the United States, second
only to mobile sources in terms of contribution
to total domestic BC. The 1.2 million tons of OC
emissions from these biomass combustion sources
represent about 75% of the total amount of OC
emitted in 2005 domestically.8
About 90% (roughly 225,000 tons) of total biomass
combustion emissions of BC in the United States
comes from "open biomass burning" sources
(Figure 4-6). Wildfires contribute about 60%
(152,000 tons) to the "biomass combustion" source
total with emissions from Alaskan wildfires alone
representing about 33% of all biomass combustion
emissions in the United States. Emissions from
wildfires can vary greatly from year to year
(Figure 4-7); however, this single year estimate of
2002 emissions is consistent with an average of
wildfire activity in the United States over the ten year
period from 2001 to 2010. About 9% (or 21,000 tons)
of the national biomass combustion total is emitted
by residential wood combustion (from "residential
heating/cooking"), and less than 2% (about 5,000
tons) from wood fired boilers and charbroiling (from
"other" sources).
7 The estimate shown applies to the total diesel PM inventory.
However, under low loads (e.g., idle), BC constitutes a smaller
fraction of PM emissions (i.e., 20-40%). Emissions in these conditions
contribute a relatively small fraction of total PM.
8 Often, global inventories define a broad "contained burning"
source category that includes the following sources from Table 4-2:
all of the sources listed in the "Residential" mega category,
"wood fired boilers" in the "Energy/Power" mega category, and
"charbroiling" in the "Other" mega category (Bond, 2007; Bond
et al., 2004). For the United States, "contained burning" sources
defined in this way em it about 27,000 tons of BC com bined (in
2005), which represents about 11% of the BC emissions and about
20% of the OC emissions from all biomass combustion (open and
contained) that occurs in the United States.
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Chapter 4
Table 4-3. National Level U.S. Emissions of PM2.5, BC, and OC for Biomass Combustion Sources in
2002/2005 (short tons). (Source: U.S. EPA)
... -r ,. i r, • Mega Source
General Specific Source Total Primary Category in
Category PM,5 Tab^4y2
Open Biomass
Burning
Residential
Heating/
Cooking
Other
Agricultural Burning
Wildfires
Prescribed Burning
Subtotal
Residential Wood
Combustion
Subtotal
Wood Fired Boiler
Charbroiling
Subtotal
Biomass Consumption Total
130,528
1,600,358
535,627
2,266,513
379,878
379,878
56,289
64,124
120,413
2,766,804
14,228
151,855
58,525
224,608
21,194
21,194
2,088
2,601
4,689
250,491
50,671
738,997
268,826
1,058,494
200,645
200,645
19,764
42,975
62,739
1,321,878
4
5
5
4.7
9.5
9.5
9
17
13.4
5
0.11
0.09
0.11
0.1
0.06
0.06
0.04
0.04
0.04
0.09
Open Biomass
Residential
Energy/Power
Other
Unlike diesel mobile sources, OC/BC ratios for
biomass combustion sources are generally much
greater than one, indicating a predominance of
OC emissions (about 80% on average). Table 4-3
further evidences a smaller OC/BC ratio (on average)
for "open biomass burning" than for the other
categories of biomass burning; however, the OC/
BC ratios are reasonably consistent at about 4 or
5 within the "open burning" categories. While the
relatively high OC/BC ratios shown in Table 4-3
for most of these sources may suggest that they
do not represent the best mitigation candidates
for climate purposes, it should be noted that OC
emissions from biomass burning may contain more
light-absorbing organic carbon ("brown" carbon)
than other sources in general (Hecobian et al., 2010;
Moosmuller et al., 2009). Exactly how much of the
inventoried OC is light-absorbing is not known at
this time.
Biomass Fired
Stationary Sources
1.9%
Residential Biomass
8.5%
Prescribed and
agricultural Burning
29.0%
Alaska Wildfires
33.0%
Non-Alaska Wildfires
27.7%
Figure 4-6. U.S. BC Emissions from all Biomass
Combustion Source Categories (250,000 short tons).
(Source: U.S. EPA)
More than half of U.S. BC emissions from wildfires
come from Alaskan fires; due to the close proximity
of these emissions to the Arctic, it is likely they
would impact the Arctic ice and snow. As discussed
in Chapter 2, both BC and OC emissions would be
expected to affect Arctic ice melt. However, as noted
above, Alaskan wildfire activity is highly variable
from year to year and peak emissions occur during
the mid-summer season, when they are less likely
to influence the Arctic due to prevailing transport
patterns during the summer. In addition, fire is a
natural ecological process in many ecosystems (see
Chapter 11).
4.3.2.3 Emissions from Energy/Power Sector
The energy/power source category contributes
approximately 7% of U.S. BC emissions and includes
a range of emissions categories, as shown in red
in Table 4-2. In general, emissions from these
sources are split fairly evenly between BC and OC.
The largest fossil fuel combustion source of BC
emissions according to the 2005 NEI is natural gas
94
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Emissions of Black Carbon
03
OJ
0>
=5
CD
U)
7,000,000 -
6,000,000
5,000,000 -
4,000,000
3,000,000 -
2,000,000 -
1,000,000
2002 2003 2004 2005 2006 2007 2008 2009 2010
Year
Figure 4-7. Acres Burned per Year in Alaskan Wildfires, 2002-2010. (Source:
U.S. Department of the Interior, Wildland Fire Management Information
(WFMI) system, https://www.nifc.blm.gov/cgi.WfmiHome.cgi}
combustion; however, estimates of the amount of
BC compared to OC in direct PM2.5 emissions from
this source category are highly uncertain.9 The
bituminous and sub-bituminous coal categories,
both of which primarily represent electricity
generating units (EGUs) but may also reflect small
contributions from commercial and institutional
sources, represent relatively small contributions to
BC emissions in the United States (a little more than
1% each). This small BC contribution is quite different
from these sources' contribution to emissions of
long-lived GHGs, where they dominate the inventory
(e.g., EGUs account for 40% of CO2 emissions).
4.3.2.4 Emissions from "Other" Source Categories
Table 4-2 shows that the remaining mega categories,
"Industrial Sources," and "Other Sources" (in blue
and white, respectively), combine to comprise
about 2% of total BC emissions domestically. As
is explained in more detail in Chapter 9, direct
PM2.5 emissions from industrial sources in the
United States are small compared to emissions of
other co-emitted pollutants such as NOX, HAPs,
and CO2. This is the result of effective control
technologies for PM emissions on a variety of
stationary/industrial sources. One industrial source
9 Specifically, EPA applies just one speciation factor to convert
direct PM2.5 emissions from natural gas combustion sources to
estimated EC emissions. This single factor is a BC/PM2.5 ratio of
0.38 which leads to a relatively large BC emissions estimate (about
25,000 tons). Though not currently available in the literature, some
unofficial source testing has suggested the BC/PM2.5 ratio is in the
range of 6 to 10% (corresponding to speciation factors of 0.06 to
0.10) indicating that both the combustion process used as well as
presence of controls on the unit will affect the amount of BC in PM2
emissions from this source type. Future work will include further
investigation into speciation for this source type.
of potential interest for additional
PM controls is stationary diesel
engines (generators, emergency
equipment, etc.), which as shown in
Table 4-2, has a low OC/BC ratio and
contributes more than half of the EC
emissions in the "Industrial Sources"
category. Existing EPA regulations
for new engines in this category are
resulting and will continue to result
in BC reductions mainly through
the use of diesel particulate filters
(DPF), although these regulations
and resultant controls do not apply
to existing engines produced before
the model year in which these
regulations became effective. Also
included in the "other" category
are many manufacturing activities
as well as fugitive dust emissions
sources and charbroiling.
4.4 Global Black Carbon Emissions
Global inventories are important for providing
information on the distribution of BC emissions
world-wide and for identifying key differences
between regions, both in terms of total quantity
of emissions and major sources. There are a few
global BC inventories available currently, and those
from Bond et al. (2004; Streets et al., 2004) are the
most widely used and referenced. Compiling a
global BC inventory is difficult for several reasons:
varying emissions among similar sources, varying
measurement techniques, different PM size cut
points used in the measurements, and the definition
of BC itself (as discussed in other parts of this
report) used in the inventories. The most up-to-
date of these inventories is for the year 2000 and
has been developed to support climate modeling
needs in the Intergovernmental Panel on Climate
Change's (IPCC) Fifth Assessment report (termed
"AR5"). These estimates have been published in the
literature (Lamarque et al., 2010) and form the basis
for all the discussion in this section. These estimates
effectively serve as "current" year, annual global BC
inventories.
In general, these global BC inventories are compiled
using fuel-consumption data to estimate emissions
from particular source categories. A few global
inventories are based on a "top-down" concept
(Parrish, 2006; Penner et al., 1993) in which emissions
are inferred from concentration and ancillary
measurements in the ambient air, usually downwind
from the source or calculated from generalized
emission factors and national or regional activity
Report to Congress on Black Carbon
95
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Chapter 4
indicators. Most global inventories, including those
of Lamarque et al. (2010) and Bond et al. (2004),
used as the basis for this section are based on
"bottom-up" type processes. EPA's inventories are
also based on "bottom-up" approaches.10 In this
method, emissions are measured or computed
directly by concentration, mass flow, and or stream
velocity observations at the source, or emissions
are calculated (using specific emission factors and
activity levels) on a source-by-source or localized
basis. Details on methods used to generate both
global and domestic BC emissions can be found in
Appendix 2, including more details on "bottom-up"
approaches.
Global BC inventories have clear advantages
when comparing emissions across world-regions,
countries, and sectors because the methodology
used is consistent across the spatial domain. Global
inventories, however, can sometimes overlook
important but subtle differences between countries
through reliance on default-type information to
estimate emissions where actual data are not
available. Regional or country-specific inventories,
on the other hand, generally contain more accurate
emissions information for the domain in question
because of the availability of more relevant and
more specific data on fuel composition, technology
differences in sectors, regulations, emission
factors, and activity levels. In this way, the relative
importance of certain sources, especially smaller
ones, can be incorporated with more accuracy
into the final emissions estimates. Unfortunately,
each regional inventory tends to employ different
methods, making comparisons across regional
inventories more difficult. Ideally, regional
inventory information could be combined with
global inventories to fill in the gaps where global
inventories are weakest. While that harmonization
has not yet fully occurred, the BC inventories
described by Lamarque et al. (2010), below, make an
attempt to combine some of the information across
global and regional inventories.
This next section provides details on global BC
inventories, including the AR5 inventory. It also
explores available regional inventories and compares
them to global inventory estimates for the same
regions. The focus of the regional comparisons will
be on Asia, where numerous regional efforts are on-
going.
4.4.1 Summary of Global Black Carbon
Emissions by Region and Source Category
Total global BC emissions for 2000 are estimated to
be about 7,600 gigagrams (about 8.4 million tons)
for 2000. The spatial distribution of these emissions
represented in Figure 4-8 shows Asia, parts of
Africa, and parts of Latin America (Central and
South America) to be among the regions emitting
the largest amounts of BC. Figure 4-9 shows global
estimates disaggregated into the these three major
world regions responsible for 75% of worldwide BC
emissions: (1) Asia (China, India, Southeastern Asia,
South Asia, Thailand, Asia-"Stan", Taiwan, Japan, and
N. Korea world regions); (2) Africa (Western Africa,
Southern Africa rest of, Eastern Africa, Northern
Africa, South Africa world regions); and (3) Latin
America (South America, Mexico, Central America,
Argentina, Venezuela, and Brazil world regions). Asia
accounts for about 40% of the global BC emissions,
Africa for about 23%, and Latin America for about
12%, as shown in Figure 4-9. Based on these AR5
emissions estimates, the United States accounts for
approximately 5% of the global total (i.e., the United
States is the 7th most significant region in the world
in terms of BC contribution).11
Table 4-4 displays total global BC emissions by 37
world regions and by 8 major source categories.
Similarly, Table 4-5 shows the distribution of the
roughly 35,700 Gg (about 39 million tons) in global
OC emissions by these same world regions and
source categories. The OC emissions from the
United States make up about 3% of the global
total.12 The last column in Table 4-4 shows the ratio
of BC emissions from each country or region to
those estimated for the United States. For example,
China (which comprises nearly all the "East Asia"
Region) emits 3.5 times as much BC as the United
States.13 It should be noted that uncertainties/
variability in the inventories (both U.S. inventories
and global inventories) could confound the ratios
presented in the last column in Tables 4-4 and 4-5.
The ratios in those tables are simply ratios of the
sum total estimate of emissions for the country
in question to the U.S. estimate (using the same
methods). Source-specific uncertainties could play a
role in the overall ranking of a country's contribution
to the total global burden of BC emissions (this
10 As an example of how these methods arrive at similar conclusions,
EPA's motor vehicle emissions model (MOVES) accurately predicts
national consumption of gasoline and diesel fuels based on vehicle
population and activity data. Differences between EPA and global
inventories may therefore be related to differences in underlying
emission rates per unit activity or fuel consumption.
11 U.S. EPA estimates of U.S. BC emissions are about 49% higher
than those from AR5, which would suggest that the U.S. actually
contributes approximately 8% to global BC em issions. These
differences are more fully discussed later in this chapter.
12 EPA estimates OC emissions at about 4% of the global total.
13 If EPA-based estimates are substituted for the AR5 estimates of
U.S. BC emissions, the ratio of China to U.S. BC emissions is closer
to 2.3.
96
Report to Congress on Black Carbon
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Emissions of Black Carbon
BC, 2000
0.1 0.2
0.5 1
10 20
Figure 4-8. Global BC Emissions based on Year 2000 Estimates, in
Gigagrams (Gg). (Courtesy of Tami Bond, produced based on data from
Bond etal., 2007)
8000
Latin America Total World
Figure 4-9. BC Emissions by Selected World Region, 2000 (Gg).
(Source: La marque et al., 2010)
of BC and OC for the various U.S.
source categories to develop a range
of emissions estimates instead of just
one number. More work along these
lines needs to be done in the future
for all emissions estimates to have a
better understanding of which sectors
vary in emissions the most and how
that variation plays a role in the
estimation of emissions.
Developed world regions like
Europe, Japan, and the Middle
East have very low BC emissions.
In these regions, like in the United
States, transportation is the
dominant emissions sector. Japan
also has significant contributions
to BC from industrial sources. In
identifying mitigation options, care
must be exercised when relying
on classifications of world regions/
countries as either "developed" or
"developing" as surrogates for BC
emissions intensity or source to
determine how "climate-beneficial"
controls might be. China, for example,
has the fastest growing economy
in the world, yet has a developing
country's level of per capita income.
While China shares the high BC
emissions levels of less developed
countries, its sources of BC are not
the same as those of less developed
areas. A crucial difference between
China and other developing areas
is China's use of coal in residential
combustion, as well as poorly
controlled emissions from industry,
and apparently a much lower reliance
on burning in agriculture than is
typical. This makes the contribution
to potential warming due to BC
emissions greater for China but
suggests the most suitable mitigation
approaches in China would be
different than in other developing
countries.
point applies to many of the graphs shown below
and is discussed further in section 4.4.4); however,
evaluation of how uncertainty plays a role in these
ratios and ranking is beyond the scope of this report.
While EPA has not done an uncertainty analysis on
its inventories, recent work by others (Chow et al.,
2011) has looked into the variability in source profiles
Figure 4-10 groups the global
emissions reported in Table 4-5 into
six broad source categories, and indicates that
global BC totals are dominated by open biomass
burning, and residential cooking and heating
sources. Roughly 35% of the total global emissions
of BC are from open biomass burning, while the
domestic (or residential) sector contributes 25%
of the global total. In developing countries, most
Report to Congress on Black Carbon
97
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vo
00
Table 4-4. Global BC Emissions in 2000 (in Gg). Transport Includes Aircraft and Shipping. (Source: Lamarque et al., 2010)
_ . r . . . T . Residential / Ag Waste ... . Grasssland Forest ... . . Country/US
Reg,on Energy Industry Transport Domestjc ^^ Waste Rres Rres Total ^
China
Western Africa
India
Brazil
Southern Africa (Rest of)
Indonesia
USA
Russia
Eastern Africa
Southeastern Asia (Rest
of)
South America (Rest of)
Australia
Western Europe (Rest of)
Central Europe (Rest of)
Japan
South Asia (Rest of)
Middle East
South Korea (Republic of
Korea)
Mexico
Northern Africa
Central America
Thailand
Canada
France
Ukraine
Argentina
Germany
Asia-"Stan"
South Africa
United Kingdom
12
0
4
1
0
1
3
5
0
1
0
0
1
3
2
0
3
3
3
0
0
0
0
1
0
0
1
0
1
1
669
20
108
53
8
28
85
33
5
30
20
11
36
26
49
13
37
55
13
11
15
20
17
10
14
12
13
10
10
10
72
15
74
91
5
34
216
32
7
45
34
12
88
40
61
30
62
36
36
36
16
33
19
48
9
26
48
2
14
31
539
127
324
30
68
73
55
102
119
101
30
7
17
54
7
68
2
9
6
37
12
12
4
11
40
6
5
27
16
4
44
8
4
3
2
12
6
7
4
3
5
4
1
2
1
0
6
3
5
1
1
2
2
0
5
7
0
2
1
0
7
3
2
2
0
1
3
1
1
1
1
0
1
1
1
1
1
1
1
1
1
0
0
0
0
0
1
0
0
0
5
505
5
70
373
7
9
35
210
6
42
120
6
2
0
1
0
0
8
0
2
3
5
0
1
14
0
25
16
0
9
105
15
215
4
252
13
145
7
166
45
20
1
3
1
2
0
0
28
0
35
12
31
0
1
4
0
0
0
0
1,358
784
538
465
460
407
390
360
353
353
178
174
150
131
123
116
111
106
99
87
84
83
78
71
71
70
68
67
58
46
3.48
2.01
1.38
1.19
1.18
1.04
1
0.92
0.9
0.9
0.46
0.45
0.39
0.34
0.32
0.3
0.29
0.27
0.26
0.22
0.21
0.21
0.2
0.18
0.18
0.18
0.17
0.17
0.15
0.12
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Energy industry Transport "-£«-' **J Waste G™ •£? Total *•»£(<*
Italy
Taiwan
Venezuela
Turkey
North Korea (Democratic
Peoples Republic)
Baltic States (Estonia,
Latvia)
New Zealand
World Total
2
1
0
1
0
0
0
54
9
18
5
12
11
1
1
1,497
31
12
7
10
1
3
3
1,340
2
2
0
2
16
11
1
1,947
0
0
1
4
0
0
0
146
0
0
0
0
0
0
0
35
1
0
8
0
0
0
0
1,481
0
0
9
0
1
1
0
1,128
46
32
30
30
29
15
6
7,628
0.12
0.08
0.08
0.08
0.07
0.04
0.01
O
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2
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Table 4-5. Global OC Emissions in 2000 (in Gg). Transport Includes Only Aircraft. Global OC Emissions Estimates Not Available for Shipping.
(Source: Lamarqueetal., 2010)
_ . r . . . T . Residential/ Ag Waste ... . Grasssland Forest ... . . Country/US
Reg,on Energy Industry Transport Domestjc ^^ Waste Rres Rres Total ^
Western Africa
Indonesia
Russia
China
Southern Africa (Rest of)
Brazil
Eastern Africa
Southeastern Asia (Rest of)
India
Australia
South America (Rest of)
USA
Canada
Mexico
Central America
South Asia (Rest of)
Central Europe (Rest of)
Asia-"Stan"
Ukraine
Thailand
Argentina
Middle East
Western Europe (Rest of)
South Africa
Venezuela
Northern Africa
South Korea (Republic of
Korea)
Japan
France
1
5
25
39
0
8
0
8
15
3
4
72
7
7
3
1
9
2
2
4
2
27
23
4
5
6
15
19
8
104
34
23
877
24
203
22
70
260
8
60
60
13
20
49
45
19
6
13
51
28
14
29
17
6
10
71
36
6
43
63
33
72
9
103
10
80
63
7
54
143
14
107
29
21
25
5
7
38
23
171
36
46
40
51
24
30
18
538
327
550
1,812
275
85
525
428
1,301
27
116
198
19
39
62
315
250
157
224
38
8
5
85
38
2
104
11
14
53
41
57
34
208
7
14
20
15
20
19
26
28
8
22
7
1
10
11
22
8
31
28
5
6
3
5
13
6
0
3
1
3
7
0
2
1
1
2
0
1
5
0
1
1
1
3
2
1
0
0
1
2
0
0
1
1
1
1
3,679
51
338
37
2,732
487
1,461
41
38
836
312
97
56
52
19
9
15
179
6
19
103
2
40
110
68
1
0
0
2
882
3,060
2,582
122
34
1,788
56
1,405
146
165
392
227
551
265
294
27
49
3
21
102
53
0
10
0
76
0
4
23
4
5,291
3,595
3,588
3,174
3,083
2,690
2,095
2,048
1,846
1,066
966
831
669
513
463
420
380
364
297
261
249
248
230
222
199
179
137
130
92
6.37
4.33
4.32
3.82
3.71
3.24
2.52
2.46
2.22
1.28
1.16
1
0.81
0.62
0.56
0.5
0.46
0.44
0.36
0.31
0.3
0.3
0.28
0.27
0.24
0.22
0.17
0.16
0.11
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-------
Region
Energy Industry Transport R«idential/ Ag Waste
a/ ' r Domestic Burning
Waste Gra«sland Forest Country/US
Fires Fires Ratio
Baltic States (Estonia, Latvia,
Turkey
North Korea (Democratic
Peoples Republic)
Germany
Taiwan
Italy
United Kingdom
New Zealand
World Total
1
8
0
11
5
9
8
0
368
1
9
13
8
29
5
6
1
2,249
2
19
1
21
12
13
14
1
1,447
64
15
28
13
1
8
7
4
7,746
0
17
0
0
0
0
0
1
696
0
0
0
1
0
1
1
0
47
1
2
0
1
0
5
1
0
10,800
11
0
16
0
1
3
0
0
12,372
79
70
58
56
46
45
36
8
35,725
0.1
0.08
0.07
0.07
0.06
0.05
0.04
0.01
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Chapter 4
of the residential (domestic)
emissions come from cook stoves
that burn biomass, dung or coal
resulting in significant emissions
of BC. China, India, and Africa to
contribute nearly two-thirds of the
total BC emissions from this source
category, an issue discussed in
more detail in Chapter 10.
Table 4-6 displays the global BC,
OC, and OC/BC ratios for 6 major
source categories. Transportation
sources show the lowest OC/BC
ratios, while burning categories
are seen to be dominated by OC
emissions and industrial sources
are somewhere in the middle.
All these sources also emit CO2
and other GHGs as well as sulfur
emissions that transform into SO2,
NOX emissions that transform into
nitrates and contribute to ozone,
and other particles.
Figure 4-11 ranks BC emissions
estimates for the 37 world regions
shown in Table 4-4, highlighting
the relative contribution of open
biomass (grassland and forest
fires) and anthropogenic sources.
With this AR5 BC inventory,
regions like Africa, Brazil, and
Australia are dominated by open
biomass burning sources whereas
countries like the United States,
China, and India are dominated by
anthropogenic sources.
Global BC Emissions, 2000 (7,600 Gg)
Other Energy/Power
0.5% 0.7%
Open Biomass Burning
(Includes Wildfires)
35.5%
Industry
19.3%
Transport/Mobile
19.0%
Residential/Domestic
25.1%
Global OC Emissions, 2000 (35,700 Gg)
Other
0.1%
Energy/Power
1.0%
Industry
6-3% Transport/Mobile
4.4%
Open Biomass Burning
(Includes Wildfires)
66.6%
Figure 4-12 details the relative
contribution of emissions for the
8 sectors in each of the 37 regions
ranked in Figure 4-11. Forest fires,
grassland fires, industry, and
transportation are all major sources
of BC depending on world region.
Areas like Asia have significant
emissions from industry, domestic,
and transportation sectors. Africa and South America
are generally dominated by open biomass burning
sources. Developed regions like the Middle East,
Japan, Europe, and the United States are dominated
by transportation sources. In the international
inventory, "nonroad" emissions are included in the
industry category, whereas in the domestic inventory
these emissions are counted in the mobile source
category. It is not possible to determine what
Residential/Domestic
21.6%
Figure 4-10. Global Distribution of BC and OC Emissions
by Major Source Category. (Source: La marque et al., 2010)
percentage of "industry" emissions are actually
"nonroad" emissions in the AR5 inventory.
Emissions estimates for BC and OC are generally
more uncertain compared to estimates for CO2, SO2,
or other pollutants primarily because BC is emitted
by a large number of small, dispersed sources with
irregular operating conditions, such as cookstoves,
biomass burning, traffic, and construction
equipment. Low technology-combustion (e.g.,
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Table 4-6. OC/BC Ratios by Broad Source Categories. (Source: Lamarque et al., 2010)
Source Category 1 BC (Gg) 1 OC (Gg) 1 OC/BC
Energy/Power
Industry
Transport/Mobile
Domestic/Residential
Open Biomass Burning
Waste
Totals
54
1,497
1,340
1,947
2,755
35
7,628
368
2,249
1,447
7,746
23,868
47
35,725
7
2
1
4
9
1
4.7
O)
^
CO
c
o
"E
LJJ
O
00
Biomass BC
Total Anthropogenic BC
Country/Region labels are
drawn directly from underlying dataset;
see source website for more details.
1000
500
i i i i 11 i i _
x^V*^ A°\^%^A°\°VVV^ o*\°VVV>
°y vy <^vvvvv ^>t>>^t
i// $^^/> ^
^ o^ ^S" ^ -C^
=/ yV
Figure 4-11. BC Emissions by World Region, 2000 (Gg). (Source: Lamarque etal., 2010, based on data from
http://acd.ucar.edu/~lamar/ipcc_ar5/bc.tar.gz)
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Chapter 4
y .*v ^ j^.^^jsk* r^.^ ..^vP ^ -&-,& v v^&'jSf'jr ^p* jS».*^>\9>-® 'r
"^
* #V» «
Forest Fires Q Waste Q Residential/Domestic Q Industry
Grassland Fires Q Ag Waste Burning Q Transport | Energy
Figure 4-12. Global BC Emissions by Source Categories and Region. (Source: Lamarque et al., 2010)
"open burning") contributes greatly to both the
emissions and uncertainties. There has not been a
lot of work done on estimating uncertainties with
BC emissions estimates. However, Bond et al. (2004)
do present a bottom-up estimate of uncertainties
in source strength by combining uncertainties in
PM emission factors, emissions characterization,
and fuel use patterns. They judge the precision of
total BC emissions estimates to be within a factor
of two. Advances in emissions characterizations
for small residential, industrial, and mobile
sources and top-down analysis combining field
measurements and modeling with iterative inventory
development will likely be required to reduce these
uncertainties further. The general "factor of 2" in
overall uncertainty estimated by Bond et al. (2004)
is comparable to the range of estimates of climate
forcing by BC given in the 4th IPCC assessment (IPCC,
2007).
4.4.2 Black Carbon Emissions North of the
40th Parallel
Emissions north of the 40th parallel are thought to
be particularly important for BC's climate-related
effects in the Arctic (Ramanathan and Carmichael,
2008; Shindell et al., 2008b). The 40th parallel north
is a circle of latitude that is 40 degrees north of
the Earth's equatorial plane. Globally, it crosses
Europe, the Mediterranean Sea, Asia, the Pacific
Ocean, North America, and the Atlantic Ocean. In
the United States, the 40th parallel approximately
bisects New York City in the East and San Francisco
in the West, passing near Trenton, NJ, Philadelphia,
PA, Columbus, OH, Indianapolis, IN, Springfield, IL
Kansas City, MO, and Denver, CO. The importance
of BC emissions, and especially marine shipping
activities (which is a significant source contributor
in the Arctic), affecting the Arctic region has been
highlighted recently in several articles and reports
(Arctic Council, 2009; Skeie et al., 2011). Arctic
shipping emissions (which are not fully characterized
in Figure 4-13) have been recently published in work
on regional inventories by Corbett et al. (2010),
Peters et al. (2011), and Paxian et al. (2010).
Global inventories indicate that most BC emissions,
particularly from fossil fuels, occur in the Northern
Hemisphere. Therefore, emissions north of
40°N latitude may be of particular concern in
understanding the impacts of BC on climate. In
addition, communities in proximity of the Arctic that
are health receptors also stand to benefit from BC
emissions reductions north of 40° latitude (these
issues are dealt with in greater detail in Chapters 2,
6, and 12). Figure 4-13 presents the magnitude of
global BC emissions and source contributions by
latitude. Transportation is the largest source of
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Emissions of Black Carbon
global BC emissions north of the 40th parallel, though
open burning, residential burning, and industrial
sources all contribute emissions north of 40°N
in the Bond et al. inventory. These patterns have
implications for assessing the contribution of source
regions to snow melt in the Arctic as well as total BC-
related forcing in the Northern Hemisphere.
While Figure 4-13 details the global distribution
of BC emissions by sector above the 40th parallel,
BC emissions from U.S. sources north of the 40th
parallel are displayed in Table 4-7. About 260,000
of the 637,000 tons (41%) are estimated to be
emitted in areas north of the 40th parallel. In terms
of the percentage of emissions from specific source
categories occurring above the 40th parallel, most
categories show BC emissions contributions north
of the 40th parallel that are proportionate to the
number of U.S. counties in that region (about 38%).
The exceptions are the Fossil Fuel Combustion and
Biomass Combustion categories. "North of 40"
emissions from biomass burning are seen to be
51% of the total domestic BC emissions from this
source category, which is attributable to the wildfire
emissions from Alaska. However, as discussed earlier
in this chapter, Alaskan wildfire activity is highly
variable from year to year, so these emissions may
vary. (Furthermore, mitigating wildfire emissions
presents particular challenges, as discussed
in Chapter 11.) BC emissions from fossil fuel
combustion north of the 40th parallel represent only
a small percentage (6%) of all emissions across the
United States for this source category.
In terms of the contribution of specific source
categories to total U.S. BC emissions (from all
North Pole
Power
Industry
Res fossil fuel
Res biofuel
Transportation
Crop waste
Open burning
South Pole
200
400
600 800 1000
BC emissions (Gg/year)
1200
1400
Figure 4-13. Geographical Distribution of Global BC Emissions by Latitude. (Source: Bond, 2008)
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Chapter 4
Table 4-7. A Comparison of BC Emissions Nationally to Those from Sources "North of 40th Parallel" in
2005 (short tons). (Source: U.S. EPA, 2005a)
Source
Total US BC
Emissions
BC Emissions
Estimated North of
the 40th Parallel
Percent of BC Emissions
From This Source
Category Emitted North of
the 40th Parallel
Percent of TOTAL BC
Emitted North of the
40th Parallel
Biomass Combustion
Fossil Fuel Combustion
Fugitive Dust Sources
Industrial Sources
Mobile Sources
Other Minor Sources
Totals/averages
250,499
43,049
1,609
6,085
333,405
2,525
637,172
128,501
2,794
483
1,574
125,784
755
259,891
51%
6%
30%
26%
38%
30%
41%
49%
1%
0%
1%
48%
0%
categories) occurring above the 40th parallel, it is
important to note that biomass burning and mobile
sources are by far the dominant contributors.
Together, these sources make up approximately 97%
of the U.S. BC emissions estimated to occur north
of the 40th parallel. The contribution from mobile
sources to total U.S. BC emissions north of the 40th
parallel is similar (48%) to the contribution of mobile
sources to total BC emissions nationally (52%), while
biomass combustion contributes 49% of total U.S.
BC emissions north of the 40th parallel (compared
to 35% of total BC emissions nationally). Again,
this reflects the heavy influence of Alaskan wildfire
emissions.
4.4.3 Alternative Estimates of Global and
Regional Emissions
In addition to the widely used Lamarque/Bond
inventory discussed above, there are other global
BC and OC emissions inventories compiled by
other researchers. Seven other global BC and OC
inventories are available in the published literature
(Cooke and Wilson, 1996; Cofala et al., 2007; Penner
et al., 1993; Junker and Liousse, 2008). The total BC
emissions estimated in these inventories fall in the
"factor of 2" error range estimated in Bond's BC
inventory, which signals that these estimates are
generally consistent with the estimates presented
above. Most of these alternative emissions are
developed using "bottom-up" approaches, similar
to that used by Bond et al. (2004) and Streets et al.
(2004). These are summarized and discussed further
in Appendix 2. The alternative emissions inventories
do not provide as much detail or as comprehensive
an explanation of uncertainty in the estimates as the
Bond inventories employed in this chapter.
An advantage of global inventories is that the
emissions estimates are compiled using consistent
definitions and methods across all regions. The
global inventories, however, do not necessarily
employ region or country specific emission factors,
activity levels, and other surrogates. Regional
emissions inventories, constructed for specific
regions, nations, or local areas, often make use of
more accurate data from local and government
sources. This may allow for improved BC emissions
estimates relative to data drawn from models or
global energy databases. Regional inventories
are more likely to account for differences in the
composition of the fuel burned, the diversity of
technologies (especially in developing countries),
and the importance of smaller sources that can
often be overlooked in global inventories. Some
of these regional inventories are based on "top-
down" type approaches while others are based on
the traditional "bottom-up" approaches described
earlier. Reconciling the global inventories with
regional inventories is complicated by differences
in methods used for each inventory. Good regional
inventories, however, may still be used to evaluate
the global estimates, and can be used to inform
future versions of those global inventories.
Most of the regional BC inventory efforts to date
have focused on the Asian sub-region (Zhang et
al., 2007; Cao et al., 2006; Sahu et al., 2008; Streets
et al., 2003a; 2003b; Ohara et al., 2007; Dickerson
et al., 2002; Mayol-Bracero et al., 2002; Reddy and
Venkataraman, 2002a; 2002b; Parashar et al., 2005)
likely due to high emissions of BC and OC from
diverse sources there. There are fewer regional
BC inventories available for European countries. In
general, global emissions inventories have to be
used to estimate European BC emissions. Recent
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Emissions of Black Carbon
2000 —
Figure 4-14. Comparison of Regional Inventories for China, India, and Indonesia with AR5 Estimates.
(Source: U.S. EPA)
work by the Arctic Council to estimate BC and OC
emissions for Arctic nations may provide useful
information on regional inventories in those nations.
A full list of available regional inventories, along
with additional details about the methods used, is
available in Appendix 2.
Figure 4-14 compares some of the different regional
BC emissions estimates for China, India, and
Indonesia to the estimates from AR5 inventories.
In general, even though the publication year of
the study (indicated on x-axis label in Figure 4-14)
is different in most cases, these inventories are
seen to be fairly consistent with one another, and
also with the Bond global inventory. The range
of emissions for a country from these various
inventories also gives an indication of the amount
of uncertainty in BC emissions estimates for a given
region. All of the regional estimates are within the
error bounds estimated by Bond et al. (2004) for
BC emissions. Recently, EPA's Office of International
and Tribal Affairs (OITA) commissioned a study
(U.S. EPA, 2011b) to look at reducing BC emissions
from various sectors in South Asia. Some of the
studies highlighted in Figure 4-14 are discussed in
more detail in Appendix A of that report, one of
the interesting things the study points out is that
BC emissions vary according to the season of the
year for certain sectors in South Asia. As a result,
emissions tend to peak during the dry season
months preceding the monsoon in South Asia.
In addition, for future work to improve the global
estimates, these regional estimates can be used
to "bound" estimates for a given world region or
country. Finally, it is important for countries to begin
developing regional inventories of BC and OC, to
better identify sources and their BC emissions, and
to supplement global inventories that sometimes
rely on "default" type information to develop
regional estimates. Having more accurate "localized"
inventories will enable better and more effectively
designed mitigation strategies for specific sources in
specific world regions.
4.4.4 Inventory Comparisons for U.S. Black
Carbon Emissions
Table 4-8 compares the U.S. portion of the 2000
global AR5-based BC and OC emissions estimates
of Lamarque/Bond et al. (in green) to the EPA's BC
estimates for 2002/2005 (in orange). U.S. emissions
from the global inventory are aggregated to the
highest level of source category detail possible
to facilitate comparisons with EPA-based BC
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Chapter 4
Table 4-8. Comparison of BC and OC Emissions (in Gg) for the United States between AR5 Global
Inventories and EPA Inventories. (Source: Lamarque et al., 2010)
BC Emissions (Gg) in AR5 and EPA Inventories
AR5 Source Description BC )lPASo"rce BC
r Description
Energy
Industry
Transport
Residential
Agricultural Waste Burning
Waste
Grassland Fires
Forest Fires
Totals:
3
85
216
55
6
3
9
13
390
Energy/Power
Industrial
Mobile Sources
Residential
Agricultural Burns
Prescribed Burns
Wildfires
39
6
302
21
13
53
138
572
EPA Estimates
High By
1200%
-93%
40%
-62%
117%
489%
962%
47%
OC Emissions (Gg) in AR5 and EPA Inventories
nncc r^ - -.- ^^ EPA Source ~-
AR5 Source Description OC Description OC
Energy
Industry
Transport
Residential
Agricultural Waste Burning
Waste
Grassland Fires
Forest Fires
Totals:
72
60
143
198
28
5
97
227
830
Energy/Power
Industrial
Mobile Sources
Residential
Agricultural Burns
Prescribed Burns
Wildfires
59
15
186
185
46
244
670
1,405
EPA Estimates
High By
-18%
-75%
30%
-6%
64%
151%
195%
69%
estimates.14 The degree of difference between
the EPA inventory and the AR5 inventory for U.S.
emissions is depicted as a percentage in the light
blue column.
Total BC emissions for the United States are
estimated to be about 390 Gg in the AR5 inventory,
14 In general, aggregating emissions from different inventories to
this level of broad source categorization introduces uncertainties
since an accurate matching of individual source categories to
these larger source categories is not always possible. The specific
source types included in the more broad categories in the AR5
inventories (and used in Table 4-8) are unclear and details were not
available for this Report. More work is needed in comparing region-
specific inventories from global estimates to regionally developed
inventories, and especially to better understand the sources that
make up the larger sectors that are generally depicted in reports
and publications.
and about 572 Gg in the EPA inventories.15 Most
of this approximately 50% difference is driven by
EPA estimates for open burning and (to a lesser
extent) for mobile sources in the United States that
are higher than those from the global inventories.
As discussed previously, wildfire emissions can vary
greatly from year to year and depend substantially
on both fire reporting systems used and emissions
calculation method (Larkin et al., 2009; 2010). This
may explain some of the difference between the
15 EPA's estimate of the domestic BC emissions in Table 4-8 (572 Gg)
is a bit smaller than the total BC emissions estimate shown earlier in
this chapter (578 Gg). This difference stems from the fact that most
of the sources in the "Other" mega source category from the U.S.
inventory were not included here. In addition, note that while an
emissions estimate (albeit small) fora "Waste" category is provided
in the global inventories, no such estimate was included in the U.S.
EPA derived inventory.
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Emissions of Black Carbon
estimates for open burning as the AR5 estimates
are based on the year 2000 and the EPA estimates
on the year 2002 and different reporting systems
and fuel loading/consumption characteristics are
used. Also, EPA estimates include all nonroad and
on-road emissions in the transportation source
category, while global inventories group emissions
from some of the smaller nonroad sources into the
"Industry" category. This could account for global
inventory estimates of U.S. emissions being lower
for "transport" and higher for "industry" compared
to the EPA estimates. In the case of OC emissions,
Table 4-8 shows that the AR5 total is about 830
Gg while the EPA estimates are seen to be about
1,405 Gg, a difference of about 69%. As with BC,
most of this discrepancy stems from fire emissions
that EPA estimated to result in more OC than do
the AR5 estimates for the United States. It is likely
that fire emissions (both OC and BC) from many
countries are under-estimated due to the methods
used to estimate fire emissions in global inventories
including an insufficient accounting for emissions
from smaller fires.
The comparison of BC emissions from the most
often used global estimates by Lamarque/Bond et al.
to BC inventories developed by EPA reveal important
differences that necessitate further investigation.
A key focus of any future examination is how these
differences may influence the estimates of regional
effects from global climate models. However, as
noted in Chapter 2, emissions uncertainties are
not thought to be as important as other factors in
determining climate impacts from model output
(Koch and Del Genio, 2010). In addition to better
understanding the role of uncertainty in emissions
estimates on impacts simulated by models, more
work is needed to better understand the source
make-up of sectors with large differences between
the two inventories (e.g., biomass combustion
sources, mobile sources, and some parts of the
residential sectors). In addition, it is necessary to
clarify the characterization of the uncertainties
associated with global BC and OC emissions
(and "the factor of 2" often discussed) estimated
by Bond et al. Development of uncertainty in
emissions estimates by sector for the U.S. and global
inventories should be a focus of future work.
4.5 Long-Range Transport of
Emissions
Aerosols emitted in a particular region can be
transported long distances through the atmosphere
to other regions of the globe. Therefore, BC emitted
in one place can affect radiative forcing in other
locations downwind. Furthermore, the climate
impacts of BC, such as effects on temperature and
precipitation, do not necessarily occur where the
radiative forcing occurs and may occur downwind
of the source region (Shindell et al., 2008b; TF
HTAP, 2010). The relationships between where
pollutants are emitted and where their impacts are
experienced are often characterized as "source-
receptor" relationships. Emissions in a source
region are transported, or lead to formation of
additional aerosols that then are transported, and
eventually deposit or affect the receptor regions
downwind. Long-range or intercontinental transport
of aerosols may occur in the planetary boundary
layer (PBL), which is the layer of the atmosphere that
is in contact with the earth's surface, or in the free
troposphere, which is the layer of the atmosphere
just above the PBL but below the stratosphere.
Aerosols that have been lofted above the boundary
layer into the free troposphere can be transported
long distances due to the relatively small amount
of precipitation and high wind speeds. In the mid-
latitudes of the Northern Hemisphere, long-range
transport is largely from west to east, due to the
prevailing winds. However, different transport
patterns are dominant in other parts of the world.
The Task Force on Hemispheric Transport of Air
Pollution (TF HTAP) organized under the Convention
on Long-range Transboundary Air Pollution
conducted a multi-model assessment of long-
range transport of aerosols and other pollutants
from four main source regions in the Northern
Hemisphere approximating the populated regions
of North America, Europe, South Asia, and East Asia
(TF HTAP, 2010). The models included in the study
produced widely varying estimates of the absolute
amount of intercontinental transport of aerosols.
Most of the diversity in model estimates appears
to be due to differences in the representation
of physical and chemical transformations that
aerosols undergo in the atmosphere, which leads to
differences in the estimated atmospheric lifetime of
aerosols. Uncertainties in emissions estimates and
atmospheric transport algorithms also contribute to
the diversity of estimates. A systematic comparison
between the TF HTAP ensemble estimates and
observations in the mid-latitudes has not been
conducted.
Although the absolute estimates in the TF
HTAP ensemble are quite different, the relative
contributions of the four continental source regions
to concentrations or deposition downwind are more
consistent. In the North American region, it was
estimated from the ensemble of simulations that
about 80% (±25%) of the BC deposited in North
America is from anthropogenic sources in North
America. Open biomass burning, largely forest
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Chapter 4
fires (wildfires), across North and Central America
contribute about 12% (±17%). Other emissions
sources from outside North America contribute
about 8% (±17%) of the BC deposited within the
North American study region.
The TF HTAP multi-model study also examined
the impact of intercontinental transport on total
atmospheric column concentrations, aerosol
optical depth, and aerosol radiative forcing. The
TF HTAP concluded that intercontinental transport
associated with anthropogenic sources of BC (not
including open biomass burning) accounted for
roughly 30% of the aerosol optical depth and
direct aerosol radiative forcing over North America.
Similarly, anthropogenic BC emissions from North
America are likely to contribute 10-30% of the BC
radiative forcing over other regions of the Northern
Hemisphere. This ensemble study would suggest that
long-range transport of BC is a minor contributor
to surface concentrations over North America, but
a major contributor to the radiative forcing and
regional climate impacts of BC. It is worth noting
that the results were calculated using rather coarse
global-scale models and variations within the North
American region were not investigated.
The results of the TF HTAP multi-model experiments
are consistent with previous modeling results that
showed that sources outside North America make
a relatively small contribution to surface aerosol
levels in North America (Chin et al., 2007; Koch et
al., 2007b) and that intercontinental transport of BC
emissions, particularly from South and East Asia,
is more important for surface concentrations or
deposition at high altitudes (Hadley et al., 2007)
and for total column loadings and climate impacts
(Reddy and Boucher, 2007; Koch et al., 2007b).
In recent work, Kopacz et al. (2011) estimated
the contribution of BC emissions sources to BC
concentrations and deposition in the Himalayas
and Tibetan Plateau and the associated direct and
snow-albedo radiative forcing. They conclude that
emissions from northern India and central China
and from western and central China contribute
most of the BC in the Himalayas and Tibetan
Plateau, respectively, although the contributions of
different locations vary with season. However, they
also show that the Himalayas and Tibetan Plateau
region can receive significant contributions from
very distant sources including biomass burning
in Africa and fossil fuel combustion in the Middle
East. They estimate that the snow-albedo effect of
BC deposition on snow in the region results in a
warming influence that is an order of magnitude
larger than the direct radiative forcing influence.
Given the paucity of anthropogenic sources of BC
in the Arctic, a large fraction of the climatic impact
of BC in the Arctic can be attributed to long-range
transport. Shindell et al. (2008b) examined the
results of the TF HTAP multi-model experiments for
insights about transport to the Arctic. Comparing
to observations of BC at Barrow, Alaska, and
Alert, Canada, all of the models appeared to
underestimate the transport of BC to the Arctic.
Consistent with the findings for the source-receptor
relationships at mid-latitudes, they found that
the models varied widely in terms of the absolute
estimates of the contribution of different source
regions, but were similar in their estimates of the
relative contributions. The ensemble results suggest
that European emissions are the largest contributors
to surface BC in the Arctic (due to the high latitude,
and therefore Arctic proximity, of many European
sources), while East Asia is the largest contributor to
BC in the upper troposphere (Figure 4-15) (Shindell
et al., 2008b). Additional source apportionment
analysis under the TF HTAP (2010) concluded that
anthropogenic emissions from Europe and open
biomass burning emissions from Eurasia both
contributed about 35% of the surface BC in the
Arctic. Anthropogenic emissions from the North
Figure 4-15. Relative Importance of Different Regions to
Annual Mean Arctic BC Concentrations at the Surface and
in the Upper Troposphere (250 hPa). Values are calculated
from simulations of the response to 20% reduction in
anthropogenic emissions of precursors from each region
(using NOX for ozone). Arrow width is proportional to the
multi-model mean percentage contribution from each
region (shaded) to the total from these four source regions.
(Source: Shindell et al., 2008b)
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Emissions of Black Carbon
American study region,
not including open
biomass burning,
accounted for an
average of 5% of
surface BC in the Arctic
region, with model
estimates spanning
the range from 2% to
10% (TF HTAP, 2010).
However, unlike the
rest of the Arctic,
deposition of BC in
Greenland, location of
the second-largest ice
sheet in the world, is
most sensitive to North
American emissions
(Shindell et al., 2008b).
In addition to the
TF HTAP approach
of largely using
grid-based models,
trajectory-based
models have also been
employed to quantify
transport to the Arctic.
These models show
a much stronger
influence of sources in Northern Eurasian locations
to Arctic surface concentrations and deposition,
and much less influence from more distant sources.
The exception is for high altitude sites in Greenland,
which may be influenced by very different sources
than the rest of the Arctic (Hirdman et al., 2010).
The contribution of both open biomass burning
and fossil fuel combustion to BC deposition in the
Arctic has been confirmed by detailed chemical
analysis of surface snow and ice cores. However,
the observational evidence would suggest that
open biomass burning, including crop burning, is
the dominant source of BC deposition in the Arctic
(McConnell et al., 2007; Hegg et al., 2010; 2009).
The relative contribution of different source types
and locations, however, varies significantly across
receptor locations and seasons.
Within the United States, the potential for transport
of domestic BC emissions to the Arctic is known to
vary by location and season. Given its proximity to
the Arctic, BC emissions sources in Alaska are likely
to have an impact on the Arctic, depending on the
synoptic weather conditions. For emissions sources
in the contiguous United States, recent trajectory
modeling work by the Joint Fire Science Program
(DeWinter et al., 2011; Larkin et al., 2011) has shown
Spring (Mar. - May)
1979 - 2009
Figure 4-16. Potential for Transport of U.S. Emissions to the Arctic based on
the percentage of spring days (March through May) in which seven-day forward
trajectories initiated within the boundary layer (below 2000m) reached the Arctic
Circle. These percentages are based on 31 years of trajectories using meteorological
data from 1979-2009. (Source: USDA Joint Fire Science Program)
that the probability of emissions impacting the
Arctic is critically dependent on the specific injection
height into the atmosphere and the specific synoptic
weather patterns prevalent at the time. Seven
day forward trajectories were computed for a 31
year period (1979-2009) using synoptic weather
patterns starting at a number of injection heights
for each location within the contiguous United
States. Figure 4-16 displays the percentage of spring
days (March-May) where any of the 7-day forward
trajectory releases below 2000m reached the Arctic
Circle. This analysis suggests that the potential for
springtime transport of BC ground-level emissions
from the contiguous United States to the Arctic
can be significant. Over the southern portion of
the United States, the potential for transport to the
Arctic is frequent (40-60%), but a significant number
of days without transport remain. For locations in
the northern part of the United States and other
higher-altitude locations, the analysis indicated that
the potential for transport trajectories to the Arctic
is very common (> 70%). However, even in areas
which show a large seasonal and climatological
potential for transport, it is possible to identify
multi-day periods where transport to the Arctic is
limited. The dependency on source location and
synoptic weather conditions may have implications
for understanding source apportionment and for
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Chapter 4
implementing mitigation strategies. Tools are being
developed to enable the use of daily predicted
transport potential in mitigation strategies.
4.6 Historical Trends in Black Carbon
Emissions
4.6.1 U.S. Black Carbon Emissions Trends
Historic trends and future projections of BC
emissions provide an indication of the relative
importance of different sectors over time and can
help focus future mitigation efforts. Establishing
emissions trends requires the use of a consistent
estimation method. Most domestic inventories
discussed earlier in this chapter are derived from
methods that have changed as measurement and
models have improved. As a result, care must be
taken in interpreting trends of over time. However,
it is possible to observe large scale changes.
Specifically, the data show that U.S. emissions of BC
increased steadily from the mid-1800s through 1920,
and then declined over the next 8 decades. This is
likely attributable to changes in fuel use from coal to
cleaner fuels, more efficient combustion of coal, and
implementation of PM controls. In more recent years,
EPA's introduction of the NAAQS for fine particles
in 1997 and strengthening of that NAAQS in 2006
necessitated PM2.5 reductions that likely contributed
to BC emissions reductions as well. In addition,
since 1990, due to regulations on PM emissions
from mobile sources, there have been substantial
reductions in BC emissions from those sources.
Since mobile source emissions are modeled, a time
series of BC emissions can be generated more easily
for this source category than for other U.S. source
categories. Mobile sources have experienced a 32%
reduction in BC, a 51% reduction in OC, and a 36%
reduction in PM2.5 emissions from 1990 - 2005. From
1990 to 2005, BC emissions decreased by 79%, 30%,
and 25% for on-road gasoline, on-road diesel, and
nonroad diesel sources, respectively. Continued
reductions are expected for mobile sources in the
next two decades, as discussed further in Chapter 8.
BC emissions for nonroad gasoline sources, though
extremely small, did not change from 1990 to 2005.
BC emissions trends for the other major source
categories (open biomass burning, industry, and
energy/power) are difficult to estimate due to lack of
data and inconsistent measurements and methods
over time. The methods used to estimate emissions
from 1990 to 2008 have changed significantly, as
has the way PM2.5 estimates are used to derive BC
emissions estimates. There are no BC estimates
available for any non-mobile source categories for
the year 1990. From 1990 to about 1998 there was
about a 30% reduction in direct PM2.5 emissions
from EGUs and other power-generation sources due
to controls on direct PM25. It is expected that some
of these reductions in direct PM2.5 led to decreases
in emissions of BC, but this is difficult to verify
without consistent speciation data for the entire
time period. In 1999, there was a major change
in the methods used estimate PM25 emissions.
Based on these new methods, from 1999 to 2008
an additional 21% reduction in direct PM2.5 is seen
from this source category. In contrast, direct PM25
emissions from industrial sources are estimated to
have declined only 6% during the entire 1990-2008
period (U.S. EPA, 20101).
Long-term trends in emissions from biomass
burning categories (wildfires, prescribed burns,
and agriculture burns) are not available due to
significant year-to-year changes in the methods
used to estimate emissions. For that reason, in the
modeling assessments "average fires" are used
to represent emissions from this source category.
However, qualitative estimates of annual wildfire
frequency/activity as well as future wildfire activity in
the United States are available and suggest upward
trends. Global climate change is expected to make
the increased activity even greater with more fuel
availability and drier, more combustion-friendly
conditions.
4.6.2 Global Black Carbon Emissions Trends
There are a number of studies available which have
looked explicitly at global BC emissions trends
over time (e.g., Bond et al., 2007; Ito and Penner,
2005; Novakov et al., 2003). Figure 4-17 (Bond et
al., 2007) shows the growth in global BC emissions
from key source categories (excluding biomass
burning) during the period between 1850 and
2000. The figure shows that emissions of BC have
increased almost linearly, totaling about 1000
Gg (approximately 1.1 million tons) in 1850, 2200
Gg (approximately 2.4 million tons) in 1900, 3000
Gg (approximately 3.3 million tons) in 1950, and
4400 Gg (approximately 4.8 million tons) in 2000.
The slower growth between 1900 and 1950 may
be due to economic circumstances and also the
introduction of cleaner technology, especially in
developed countries. OC shows a similar pattern of
linear growth that is slightly slower in the mid-1900s.
Figure 4-18 relates BC emissions trends from Bond
et al. (2007) to earlier work done by Ito and Penner
(2005), and by Novakov et al. (2003).16 Ito and
' Novakov et al. looked at BC from fossil fuel combustion only.
7 72 Report to Congress on Black Carbon
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Emissions of Black Carbon
• Aviation fuel
S Light distillate
Q Middle distillate
n Residual oil
E3 Coal
D Biofuel
K
1850 1875 1900 1925 1950 1975 2000
1850 1875 1900 1925 1950 1975 2000
B Latin America
H Middle East
ffl India
CD Africa
D Other Asia/Pacific
• Europe
E3 Former USSR
H China
Q North America
1850 1875 1900 1925 1950 1975 2000
1850 1875 1900 1925 1950 1975 2000
Figure 4-17. Historical Growth in Emissions of BC (Panels a, c) and OC (Panels b, d). Segregated by Fuel
(Panels a, b) and World Region (Panels c, d). (Source: Bond et al., 2007)
7000
1850 1875 1900 1925 1950 1975 2000
6000
10000
8000 •]
1850 1875 1900 1925 1950 1975 2000
12000
(0
o 1000
0
1850 1875 1900 1925 1950 1975 2000
- Ito and Rentier
- This work
1850 1875 1900 1925 1950 1975 2000
Figure 4-18. Historical Reconstruction of Global Emissions Trends. Comparison of Bondetal. (2007)
("this work") with Previous Studies.
Report to Congress on Black Carbon
113
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Chapter 4
1.4
1.2
>• 1.0
D)
0.8
m
0.4
0.2
0.0
a) United States
•>-.*.
A A A A A A A
• . • • o
0.3
0.2
0.1
0.0
b) United Kingdom
2.5
2.0
1.5
c) China
O residential/commercial
• industry
A diesel
n utilities
« total
1900 1920 1940 1960 1980 2000 1950 1960 1970 1980 1990 2000 1950 1960 1970 1980 1990 2000
Year
Figure 4-19. BC Emissions (Tg /y) in the United States, United Kingdom, and China (Novakov et al.,
2003). BC emissions are estimated from annual consumption data for the principal BC producing fossil
fuels and BC emission factors disaggregated by utilization sector. BC from biofuels and open biomass
burning are not included.
Penner show a very similar trend and magnitude in
BC emissions from biofuel, but the magnitude of
fossil-fuel BC emissions is much lower. In the late
1900s, Bond et al.'s biofuel emissions increase less
(about 30% between 1960 and 2000 vs. 100% for Ito
and Penner). By contrast, Novakov et al. estimated
higher fossil-fuel BC emissions than Bond et al., in
the early 1900s. Novakov et al.'s work was based
on total BC aerosol, while Bond et al. and Ito and
Penner's work focused on fractions of PM less than
1 micron in diameter. Novakov et al. shows flat BC
emissions between about 1910 and 1950, similar to
Bond et al. (2007).
The greatest difference between the more recent
Bond et al. (2007) work and the earlier Ito and
Penner and Novakov et al. work is the more gradual
transition in the latter half of the 20th century. Both
of the earlier studies considered the introduction
of cleaner diesels and some changes in sectoral
divisions. Bond et al. modeled shifts to cleaner
burning through increases in consumption in cleaner
sectors. Bond et al. indicate the shift to cleaner
burning coal explicitly for the first time, reducing
BC emissions from this sector. It is likely that the
difference between the three studies is largely
attributable to the choice of emission factors, which
entails some implicit assumptions about technology
choices.
Figure 4-19 shows the estimated BC emissions
trends for the U.K., United States, and China
(Novakov et al., 2003). According to these data,
emissions from the United States peaked in 1920,
while Europe peaked in 1950 and has declined
about 90%. Total global emissions of BC, however,
have been steadily increasing since 1875 (Novakov
et al., 2003). Presently, global BC emissions total
approximately 8.4 million tons. Almost all of the
increase in recent decades is from developing
countries in Asia, Latin America, and Africa. China
and India contribute nearly 25% of global BC
emissions.
Together, these emissions trends studies and
other works suggest that developed countries
dominated global BC emissions until the adoption
of pollution control technologies and fuel-use shifts
began to slow growth and eventually to result in
significant reductions after mid-century (Bachmann,
2007; 2009; Ramanathan et al., 2007). Available
data suggest that BC emissions from developed
countries have declined substantially over the past
several decades, while emissions from developing
countries have been growing. Today, the majority of
BC emissions are from developing countries (Bond
et al., 2007) and this trend is expected to increase
(Jacobson and Streets, 2009; Kupiainen and Klimont,
2007).
774
Report to Congress on Black Carbon
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Chapter 5
Observational Data for Black
Carbon
5.1 Summary of Key Messages
• Estimates of BC are made with a variety of
instrumentation and measurement techniques.
Most ground level estimates of BC are reported
as mass concentrations based on thermal-optical
(EC) and filter-based optical (BC) techniques. BC
and EC values from these measurement methods
are highly correlated, although the method-
defined values may differ by as much as a factor
of two; however, consistent measurements in
long-term monitoring networks are sufficient
to detect trends that correlate with emissions
reductions. Published studies show that BCEC
ratios derived by commercial instrumentation are
generally within 30%.
• The United States has recently standardized BC
measurements for its major routine speciation
monitoring networks. Additional research is
needed to further standardize ambient and
emissions measurement methods and to develop
factors that harmonize existing measurements
produced from different sampling and analytical
techniques. It is also recommended that light
absorption be reported in the original units
of absorption along with any mass absorption
coefficients or conversion factors used to convert
absorption to BC mass concentration.
• Ground-level BC measurements across the globe
indicate estimated concentrations ranging from
< 0.1 ug/m3 in remote locations to ~15 ug/m3
in urban centers. Although monitor locations
are sparse globally, available observations
suggest ambient levels in China are almost 10
times higher in urban and rural areas than those
in North America or Europe. A comparison of
urban concentrations to corresponding regional
background levels reveals an urban increment
of up to 2 ug/m3 in North America and Europe
compared to an urban increment of -6-11 ug/m3
in China.
• In the United States, BC comprises -5-10% of
average urban PM2.5 mass.
Long-term records of historical BC
concentrations, derived from sediments or ice
cores, valuably supplement available ambient
data.
- Long term trends in estimated ambient
concentrations derived from BC in
sediments of the New York Adirondacks
and Lake Michigan show recent maximum
concentrations occurred in the early- to
mid-1900s and it appears concentrations
have since decreased, which is attributed to
decreased U.S. fossil fuel BC emissions.
- Ice core measurements in Greenland reveal a
similar maximum BC level in the early 1900s
related to industrial emissions, but also show
that biomass burning emissions contribute
significantly to deposited BC in the Arctic.
- Globally, Northern Hemispheric ice core BC
trends vary with location; some ice cores have
BC values increasing to present-day, while
other areas show maximum levels reached
earlier in the 1900s.
Over the past two decades when U.S. ground-
level ambient BC measurements are available,
ambient BC concentrations have declined
substantially, most likely due to reductions in
mobile source emissions and other controls on
direct PM2.5 emissions. Since 2007, the decline
may be due in part to recession-related decreases
in vehicular travel and industrial output. The
ambient concentration declines appear to be
stronger in urban areas and may in fact be larger
than the estimated average reductions in total BC
and direct PM2.5 emissions in the United States.
Estimates of the total atmospheric column using
remote sensing qualitatively show similar spatial
variability in absorbing aerosol levels across the
globe to ground level measurements. Remote
sensing measurements that utilize multiple
wavelengths also show that the absorbing
particle mixture varies globally among areas
dominated by urban-industrial sources, biomass
burning, and wind-blown dust.
Report to Congress on Black Carbon
115
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Chapter 5
5.2 BC and Other Light-Absorbing
Carbon: Measurement Methods
Current measurement techniques generally estimate
BC on the basis of light absorption characteristics
or by thermally isolating a specific carbon fraction.
The techniques used currently to estimate BC
mass concentrations are summarized in Table 5-1.
These two general categories of BC measurement
techniques can be viewed as different indicators of
the chemical and physical properties of BC.1 This
is discussed further in the text box, "Measurement
Approaches for BC." The two most common BC
measurement techniques are thermal-optical and
filter-based light absorption as denoted in the table
below.
for pyrolysis or the charring of organic material
during analysis. Despite the seeming simplicity
of thermally separating particulate carbon into
two fractions, there is considerable uncertainty in
assigning carbon mass to either OC or EC fractions.
For example, charring of particles during the
thermal analysis has produced erroneous OC and
EC assignments (Cadie et al., 1980; Huntzicker et
al., 1982; Yu et al., 2002b). In addition, there are
several different commonly used temperature
protocols that cause variation in the OC and EC
assignments. Watson et al. (2005) and Chow et al.
(2006) provide a detailed summary of the variety of
methods used that demonstrates the wide range of
thermal analysis protocols in use. The two thermal-
optical methods that are predominantly used in
Method Type
Light absorption/optical
Table 5-1. Description of BC Measurement Techniques.
Method Description
Filter-based: Light absorption by particles is measured through a filter loaded with
particles; BC is quantified using factors that relate light absorption to a mass concentration.
Photoacoustic: Light absorption by particles is measured by heated particles transferring
energy to the surrounding air and generating sound waves; BC is quantified using factors
that relate light absorption to a mass concentration.
Incandescence: Incandescent (glowing) particle mass is measured; BC is quantified by
calibrating the incandescent signal to laboratory-generated soot.
Prevalence
of Use
High
Low
Low
Isolation of specific
carbon fraction
Thermal-Optical: BC is measured as the carbon fraction that resists removal through
heating to high temperatures and has a laser correction for carbon that chars during the
analysis procedure; BC is quantified as the amount of carbon mass evolved during heating.
Thermal: BC is measured as the carbon fraction that resists removal through heating to
high temperatures; BC is quantified as the amount of carbon mass evolved during heating.
High
Low
Thermal-optical measurements involve exposing
a particle-laden filter to a series of heating
steps. These measurements involve a multi-step
temperature program to evolve OC in pure helium
and EC in a helium/oxygen atmosphere with an
optical (transmittance or reflectance) correction
1 In current practice, measurements produced from light absorption/
optical methods are expressed as BC while those produced from
thermal-optical or thermal methods are referred to as EC. To
simplify the discussion, this differentiation in characterization of
BC by measurement method is not repeated. Instead, since both
measurement types are essentially estimating the same parameter
(i.e., BC) albeit via different method orientation, and to make clear
that light absorption measurements do not necessarily provide a
'better' indicator of BC than thermal methods, the term BC is used
to describe all measurements. In Appendix 1, where this topic is
more thoroughly explored, the BC measurements produced by light
absorption/optical methods are referred to as apparent BC or "BCa",
and those produced by thermal or thermal-optical methods are
referred to as apparent EC or "ECa".
monitoring networks in the United States are the
thermal-optical transmittance (TOT) and thermal-
optical reflectance (TOR) methods (Chow et al., 1993;
Chow et al., 2007; Peterson and Richards, 2002).
The TOT method differs from the TOR method
in the thermal combustion program used and
the method of correcting for char (transmittance
versus reflectance). Long-standing reliance on the
thermal optical methods has resulted in an extensive
observational record based on OC/EC splits, and the
frequent use of EC as an estimate of BC.
While EC is directly quantified as the mass of
carbon atoms that evolve during a thermal or
thermal-optical analysis, optical techniques observe
the light-absorbing properties of the particles to
estimate BC. Filter-based, optical instruments are
relatively low cost, readily available, and simple to
operate, and thus are frequently field deployed
776
Report to Congress on Black Carbon
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Observational Data for Black Carbon
Measurement Approaches for BC
The chemical and physical properties of carbonaceous PM vary in terms of both refractivity (the inertness of the carbon at
high temperatures) and light absorption. Each carbon measurement technique provides unique information about these
properties. All current analysis methods are operationally defined, meaning that there is no universally accepted standard
measurement approach. When developing methods and operational criteria to identify BC, some scientists use its optical
properties or light-absorbing characteristics (optical or light absorption methods), some use its thermal and chemical
stability (thermal-optical methods), while others use its morphology or microstructure or nanostructure (microscopy
methods). One major class of methods, thermal or thermal-optical techniques, distinguishes refractory and non-refractory
carbon as EC or apparent EC (ECa), and OC or apparent OC (OCa) respectively (see Figure 5-1).The second major class of
methods, optical methods, quantifies the light absorbing component of particles as BC or apparent BC (BCa), which can
be used to estimate BC concentrations and can also indicate the existence of components that absorb in the near-UV (i.e.,
brown carbon, BrC). Light absorbing carbon (LAC) is a term used for light-absorbing particles in the atmosphere, which
includes BCand BrC.
Light-Absorption Classification
More
light-absorbing
Light-
absorbing
carbon
(LAC)
Black
carbon
Brown
carbon
(BrC)
Less
light-absorbing
Thermal-Optical Classification
More refractory
Elemental
carbon
Organic
carbon
(OCJ
Less refractory
* Measurementtechnique-specificsplitpoint
Figure 5-1. Measurement of the Carbonaceous Components of Particles. Black carbon and other types
of light-absorbing materials can be characterized by measuring their specific light-absorbing properties,
as seen on the left side of the figure (BCa/BrC/LAC). This contrasts with other approaches to characterizing
particles based on measurements of the refractory nature of the material (inertness at high temperatures),
as seen on the right side of the figure (ECa and OCa). (Source: U.S. EPA)
to measure BC. Filter-based instruments measure
the quantity of light transmitted through a filter
loaded with particles (Hansen et al., 1982; Lin et al.,
1973; Rosen and Novakov, 1983). For filter-based
optical instruments, the detected light absorption by
particles can be converted to an estimated BC mass
concentration. There are two main uncertainties
associated with the quantification of filter based BC
using optical methods: 1) a filter loading artifact
and 2) the selection of an appropriate conversion
factor. Several studies have shown that filter-based
BC measurement can be affected by the amount
and composition of particles loaded onto the filter
(Arnott et al., 2005; Collaud Coen et al., 2010; Park
et al., 2010; Schmidt et al., 2006; Virkkula et al.,
2007; Weingartner et al., 2003). In addition, the
selection of the conversion factor to relate light
absorption to mass is a significant issue. No single
factor is applicable to all methods, wavelengths,
particle sizes, particle compositions, shapes and
structures. There are a variety of conversion factors
that have been published in scientific literature and
are commonly applied to estimate BC (Gundel et al.,
1984; Liousse et al., 1993; Petzold et al., 1997; Bond
and Bergstrom, 2006; Novakov, 1982). Theoretical
and empirical studies show that bounds can be
placed on absorption efficiencies for different
assumptions of BC aerosol origins and composition
Report to Congress on Black Carbon
117
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Chapter 5
(Alfaro et al., 2004; Andreae et al., 2008; Dillner et
al., 2001; Cross et al., 2010; Chou et al., 2005; 2009;
Favez et al., 2009; Fu and Sun, 2006; Fuller et al.,
1999; Horvath, 1993; Jacobson, 2010; Knox et al.,
2009; Liousse et al., 1993; McMeeking et al., 2005;
Nordmann et al., 2009; Ogren et al., 2001; Ram
and Sarin, 2009; Ramana et al., 2010; Rosen and
Novakov, 1983; Schuster et al., 2005; Subramanian et
al., 2010; Watson et al., 2005; Widmann et al., 2005;
Chan et al., 2010). A suggested solution would be
to quantify BC in the original light absorption units,
which is a strength of the optical techniques. It is
recommended that light absorption be reported in
the original units of absorption along with any mass
absorption coefficients or conversion factors used to
convert absorption to BC mass concentration.
While the terms "BC" and "EC" are frequently
associated with measurements from the two general
categories of specific commercial instruments in
the scientific literature, both of these measurement
techniques provide estimates of BC concentrations
(Wolff et al., 1982; Andreae and Gelencser, 2006).
Ambient monitoring studies that simultaneously
utilized light absorption and thermal-optical
methods show that the estimates of BC by the two
techniques are on average near 1 and generally
within 30% (70% of studies had ratios between 0.7
and 1.3); however there do exist studies reporting
very low BCEC ratios (-0.5) and very high BCEC
ratios (~2). Ambient inter-comparison studies
have found that estimates of BC from thermal
measurement methods are usually reliable predictors
of ambient BC estimated via light absorption
techniques and vice versa. The comparison of
EC by thermal-optical methods and BC by light
absorption is sensitive to the source of EC/BC and
varies by location. While the estimates from the
two techniques are highly correlated and display
similar concentration values, they can vary by up to
a factor of two among the limited number of studies
available.2 Further discussion of these comparisons
can be found in Appendix 1.
supplement these thermal measurements. There is
also a modest network of BC monitoring sites across
the globe in remote areas to provide information
about background levels.
5.3.1 Major Ambient Monitoring Networks
Figure 5-2 provides a map showing the extent of
known BC monitoring networks around the globe.
The existing networks in the United States, Canada,
Europe (EUSAAR, EMEP), and Asia (CAWNET), as well
as those with global coverage (GAW, ESRL/GMD)
and ad hoc collections of special study data are
shown.3 The map separately shows locations using
light absorption, thermal, or both measurement
techniques. Most locations shown are in North
America and these monitors mostly utilize thermal
measurement techniques.
Ambient BC data in the United States are mostly
available from PM2.5 urban and rural speciation
monitoring networks which use thermal
measurements. The Interagency Monitoring of
Protected Visual Environments (IMPROVE) network
started collecting data in 1987, and the urban
Chemical Speciation Network (CSN) started in 2001.4
Urban BC is measured through the CSN network of
approximately 200 monitors located in major urban
areas.5 In rural environments such as national parks
and wilderness areas, the United States relies on the
IMPROVE network to characterize air quality. This
network consists of approximately 160 monitors.
Like the CSN, the IMPROVE network utilizes thermal
measurement technologies.6 Other U.S. data include
supplementary measurements from approximately
45 semi-continuous light absorption monitors
(operational in 2007); 5 semi-continuous carbon
measurements; and smaller networks of thermal-
optical and light absorption monitors (SEARCH,
Super-sites). See Appendix 1 for more details.
Thermal-optical methods for measuring OC and EC
are currently standardized and consistent among
the IMPROVE (http://vista.dra.colostate.edu/improve),
5.3 Ambient Concentrations of BC
Currently, few countries have robust networks for
ambient measurement of PM2.5. Most available
global ambient BC data are produced in the United
States, Canada, Europe, and China, and the vast
majority of these data are based on the more widely
available thermal measurement techniques (see
section 5.2). In the United States and Europe, limited
light absorption measurements are available to
2 Comparable studies of the relationship between measured
estimates of BC from light absorption and thermal techniques have
not been conducted for direct measurements of source emissions.
3 These BC measurements are publicly available or have been
included in peer-reviewed publications.
4 The VIEWS web site (http://vista.dra.coiostate.edu/views/') provides
information on the start and end dates for each site.
5 Measurements are based on integrated 24-hr samples, mostly
collected every three days, and were mostly analyzed for EC
between 2001-2007 using an EPA NIOSH-type TOT protocol. EPA
started to transition CSN measurements to the IMPROVE_A TOR
protocol for EC in May 2007.
6 Measured every three days. The IMPROVE program slightly
modified the protocol in 2005, which resulted in higher quality
data and slightly higher EC as a fraction of total measured carbon.
The IMPROVE network data for 2005-2007 are produced using the
newer IMPROVE_A TOR protocol.
778 Report to Congress on Black Carbon
-------
Observational Data for Black Carbon
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Figure 5-2. Ambient BC Measurement Locations Worldwide. Light absorption measurement locations
are colored black. Thermal measurement locations are colored in red. A small subset of locations with both
measurements is colored yellow. (Source: U.S. EPA)
CSN (http://www.epa.gov/ttn/amtic/speciepg.html),
and SEARCH (http://www.atmospheric-research.com)
networks.
5.3.2 Global Ambient Concentrations
Table 5-2 summarizes data from a number of studies
and monitoring networks that help illustrate the
range of BC concentrations across the globe. The
table also indicates the BC measurement methods
(thermal (T) and light absorption (LA)) for each
study/monitoring network. While BC measurements
for urban and rural areas are similar in North
America and Europe, the reported concentrations
for China are much higher. Both urban and rural BC
concentrations in China are approximately 10 times
higher than urban and rural concentrations in the
United States, respectively.7
The United Kingdom shows higher BC concentrations
at the upper range than the United States likely due
to the influence of local sources on the individual
monitoring sites. In general, roadside or near-source
7 As discussed in Chapter 4, the ratio between China and U.S.
measured BC concentration is two to three times higher than the
ratio of their estimated national BC emissions.
monitors yield higher values, as demonstrated
by the curbside monitors in London which report
considerably more BC than the urban-wide locations
(Butterfield et al., 2010). The "Black Smoke" data for
the UK that provide the basis for the five-decade
trend discussed in section 4.6.2 are three to four
times higher than co-located measurements of BC
(Quincey, 2007).
The global background sites that are part of the
National Oceanic and Atmospheric Administration
(NOAA) network reveal BC concentrations that
are one to two orders of magnitude lower than
those typically observed in either urban or rural
continental locations. The presence of BC in these
remote locations without any nearby sources is
indicative of long range transport and is used to
evaluate intercontinental transport processes in
global models.
5.3.3 Comparison of Urban and Rural
Concentrations Globally
Available data suggest that BC concentrations
vary substantially between urban and rural areas.
Specifically, urban areas tend to have higher
Report to Congress on Black Carbon
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Chapter 5
Table 5-2. Summary of Selected Global BC Ambient Concentrations for Urban and Rural/Remote Areas.
Range of Annual Average Concentrations (ug/m3)
Region Networks Year Method Type ^ *™^**
United States
Canada
Europe
Europe
United Kingdom
China
Nepal
Global
Background
Other Arctic
Sites
United Kingdom
CSNV IMPROVE"
SLAMSC
NAPSd
EMEPe
EUSAAR'
BC Network9
CAWNET11
NCO-P1
NOAAGMD Sites'
Mauna Loa
Point Barrow
South Pole
Alert (Canada)
Zeppelinfjell'
(Svalbard, Norway)
Black Smoke (BS)k
2005-2007
2007
2003-2009
2002-2003
2006
2009
2006
2006-2008
1990-2006
1988-2007
1987-1990
1989-2008
2002-2009
2006
T
LA
T
T
T
LA
LA
T
LA
LA
LA
LA
LA
LA
LA
0.3 to 2.5 (-200 sites)
0.3 to 3.0 (~ 45 sites)
0.9 -1.8 (12 sites)
1.4 -1.8 (2 sites)
1.5 (2 sites)
2.7 (1 site)
1.0 -2.9 (19 sites)
9.3 -14.2 (5 sites)
5.0 -16.0 (12 sites)
0.1 -0.6 (-150 sites)
0.4 - 0.8 (4 sites)
0.2 -1.8 (12 sites)
0.1 - 0.7 (4 sites)
0.2 - 0.5 (4 sites)
0.3 -5.3 (13 sites)
0.16(1 site)
0.01 -0.02
0.02 - 0.07
0.002-0.004
0.04 - 0.1
0.02 - 0.06
i CSN - Primarily urban network sites.
. IMPROVE - Rural network sites.
BC data at State and Local Air Monitoring Stations from AQS, mostly with Magee Aethalometers.
i Personal communication with Tom Dann (Environment Canada).
• Monitoring was for the period 07/02 - 06/03 from Yttri et al. (2007).
Data taken from http://ebas.nilu.no/or EUSAAR, the sites assigned to be urban are Ispra, IT (BC) and Melpitz, DE.
Although not part of EUSAAR, the urban sites also include Ring A10, NL (EC). The northern EUSAAR remote location of
Zeppelinfjell, NO, site is included with other Arctic sites listed separately.
i Urban network sites from Butterfield (2010); Curbside site at London Marylebone Road reported 10ug/m3.
i Data and urban/regional/remote classification was for the period 2006 from Zhang et al. (2008).
Monitoring was for the period 03/06 - 02/08 from Marinoni et al. (2010).
NOAA Global Monitoring Division Sites - For this table, we modified reported numbers in absorption units using a
nominal mass extinction coefficient of 10m2g-l. One year from each site was eliminated as non-representative.
: Data taken from http://www.airquality.co.uk/reports/cat05/1009031405_2009_BC_Annual_Report_Final.pdf; curbside site
at London Marylebone Road reported an average of ~40ug/m3 for each year.
concentrations. The global BC data (for 2005-
2007 average or calendar year 2006) displayed in
Figure 5-3 contrast the annual average rural and
urban concentrations for North America, China, and
Europe.8'9The ambient rural concentrations provide
an indicator of regional background concentrations
8 The data in Figure 5-3 are aggregated and displayed on the 1.9 x
1.9 degree resolution which is widely used by global climate models.
This coarse grid does not allow us to see sharp gradients which tend
to exist within urban areas. Also, note that these grid-based displays
use a logarithm ic scale to show the order of magnitude range of
concentrations for BC across the globe.
9 The map in Figure 5-3 shows the 40th parallel, the importance of
which is discussed further in Chapter 4.
resulting from regional emissions and transported
aerosols. Levels in urban areas reflect the higher
average concentrations resulting from the
combination of local emissions and regional
emissions. The portion of urban concentrations
due to local emissions can also be described as the
"urban increment" or "urban excess".10
10 Because of strong regional homogeneity among background
measurements, urban grid squares without measurements were
estimated from nearby cells to permit an estimate of urban excess.
These estimated values may be higher than surrounding regional
measurements. Spatial interpolation here is based on inverse
distance weighting of the nearest neighbors (Abt Associates, 2005).
720
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Rural
Urban
Observational Data for Black Carbon
Urban Excess
United States and Canada
0.2-1.9ng/m3
ff
, 3 <-^
g/mJ f\
. 3 -
..
—• *>
14.2
Figure 5-3. Spatial Distribution of Global BC Data. Rural, urban, and urban excess concentrations
for 2005-2007. Grid squares with a white dot represent estimated rural concentrations from spatial
interpolation of the nearest neighbors with measurement data. The 40th parallel is shown as a dotted line.
(Source: U.S. EPA)
As demonstrated in Figure 5-3, urban BC
measurements in North America are generally
much higher than the nearby regional background
levels. This suggests that there can be a substantial
increment of local emissions in urban areas. For the
period 2005 to 2007, the urban increment ranged
from zero to 2.2 u.g/m3 (i.e., up to 92% of the total
urban BC concentrations). In general, average urban
concentrations are relatively consistent across North
America, though the larger populated regions of the
eastern United States, eastern Canada, and California
contain most of the highest concentrations. However,
the western United States and western Canada have
lower regional background concentrations and
therefore relatively larger urban increments, while
higher rural concentrations in eastern North America
result in smaller urban increments (more similar
regional and urban average values). The higher
regional background levels across eastern North
America suggest higher and more consistent levels
of BC emissions from sources across the region,
and/or greater transport from clustered cities to
surrounding rural areas.
Figure 5-3 also shows that Europe's measurement
data are quite similar to those for North America.
However, both China's regional and urban BC
concentrations are much higher than those seen in
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Chapter 5
• IMPROVE Site
» IMPROVE Urban Site
A CSNSite
Figure 5-4. Annual Mean BC Concentrations (ug m3) for 2005-2008 in the United States. Panel (a)
shows rural EC concentrations provided by IMPROVE network. Panel (b) shows the combined rural
IMPROVE and urban CSN network EC data. IMPROVE site locations are shown as black circles, CSN sites
are shown as black triangles, and urban IMPROVE sites are shown as magenta diamonds. (Adapted from
Figures 7.5.1 and 7.5.2 of IMPROVE report V (2011) http://vista.cim.colostate.edu/improve/)
North America and Europe and its urban increments
are approximately four times larger. This can be
attributed in part to larger urban and regional
emissions sources in China compared to North
America and Europe.
To provide additional spatial detail to U.S. BC
concentrations, Figure 5-4 shows the spatial
distribution of background and combined urban
plus background concentrations derived from EC
measurements in the IMPROVE and CSN monitoring
programs. While the background concentrations
are relatively homogeneous over large sub-regions
of the United States, Figure 5-4(a) indicates these
background concentrations are generally higher in
the Industrial Midwest. As shown in Figure 5-4(b),
urban areas have much higher concentrations.
High ratios of urban to rural BC concentrations
722
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Observational Data for Black Carbon
BC (absorption units)
<0.8
Community Districts
Figure 5-5. Urban BC Gradients for New York City.
(Source: The New York City Community Air Survey, Results
from Winter Monitoring 2008-2009, http://www.nyc.gov/)
demonstrate the localized impact of BC on
surrounding rural regions and suggest urban BC
emissions sources were significantly larger than rural
sources. The mean urban to rural ratio was 3.3 ± 1.9
and was much larger than the mean ratio for sulfate,
nitrate, or OM components of PM25 (Hand et al.,
2011).
In addition to the general contrast between
measured urban and nearby rural BC concentrations,
there can be substantial spatial variation in BC
concentrations within a given city. Because global
representations of BC concentrations are typically
based on limited monitoring locations and are
generally presented as average concentrations
(often across monitors hundreds of miles apart), it is
important to realize that ambient concentrations of
BC in any urban area can vary widely from location to
location within the city. BC concentrations can vary
spatially within an urban area because the magnitude
of monitored BC concentrations is dependent on
the proximity of the monitor to roadways and other
nearby sources. Therefore, concentrations measured
may not be representative of other locations.
Figure 5-5 illustrates the estimated spatial variability
of BC in New York City.11 This special study used
150 monitoring sites to reveal large gradients in
apparent BC concentrations. Actual gradients may
even be larger. Furthermore, the illustrative within-
urban variability for NYC may also be representative
of other urban areas with high population and
emissions density. While most of the identified high
concentration zones can be attributed to mobile
source emissions density, this study also revealed
significant BC emissions sources associated with
residential oil combustion.
5.3.4 BC as a Percentage of Measured
Ambient PM2 5 Concentrations in the United
States
Because total PM2.5 mass is the basis for regulation
of fine particles in the United States and also
serves as the basis for BC emissions estimates, it
is informative to estimate the contribution to total
PM2.5 mass from BC. However, given the limited BC
data available on a global scale, this evaluation is
based solely on data for urban areas in the United
States that are regionally representative of large
U.S. cities. Compared to U.S. rural locations, urban
locations contain a higher percentage of BC and
OC. While urban nitrate concentrations are also
higher than surrounding rural areas, carbonaceous
aerosols are responsible for most of the urban
PM2.5 increment. Other components, such as dust,
are similar in both urban and rural environments
(U.S. EPA, 2004b). Figure 5-6 shows the BC fraction
of PM2.5 mass for 15 selected U.S. urban areas. The
values represent average concentrations among
monitoring locations in the area. The average BC
concentrations range from 0.6 u.g/m3 in St. Louis to
1.2 u.g/m3 in Atlanta. The percentage of PM2.5 that is
BC ranges from 4% in St. Louis to 11% in Seattle.12
A more complete characterization of urban and
rural PM25 speciation components on an annual and
seasonal basis can be found elsewhere (Hand et al.,
2011; U.S. EPA, 2009b).
11 Based on 150 filter-based portable sam piers and optical
absorption measurements with the smoke stain reflectometer.
12 Approximately 20-80% of the estimated ambient organic matter
(OM) is directly emitted (Carlton et al., 2009). The other portion,
termed secondary organic aerosol (SOA), is formed through
chemical reactions of precursor emissions after being released
from the sources (Saylor et al., 2006; Carlton et al., 2009; Chu, 2005).
OM is typically 1.4 to 1.8 times higher than measured OC levels in
urban areas, with an even larger multiplier of OC levels measured
in rural areas (Bae et al., 2006; Turpin and Lim, 2001). The OM-to-
OC ratio tends to be higher with an aged aerosol (resulting from
transported, atmospheric-processed, and aged particles), SOA, or
directly emitted OM from biomass combustion. Although we are
not able to quantify the amount of OM that may be BrC, it is worth
noting that average OM for the 15 selected cites represents 26% to
55% of PM2.5 and the OM-to-BC ratio ranges from 4 to 9.
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Chapter 5
Figure 5-6. Composition of PM2.5
for 15 Selected Urban Areas in the
United States. Annual average PM2.5
concentrations (ug/m3) are presented
where the circle size represents the
magnitude of PM2.5 mass. The BC
and Organic Mass (OM) fractions are
illustrated. OM represents OC together
with its associated non-carbonaceous
mass (e.g., hydrogen, oxygen and
nitrogen), estimated by a material
balance approach. Sulfates and Nitrates
have been adjusted to represent their
mass in measured PM2.5. (Source: U.S.
EPA, 2009b)
1 »JO
o
i
£
V
u
e ,,,
i
3 KiO
C
Start of the National Survey
Year (Apr-Mar up to 1961-62, then calendar years 1962 - 2005}
Figure 5-7a. Trends in Black
Smoke Measurements
(ug/m3) in the United
Kingdom, 1954-2005. The
BS measurements are highly
correlated with optical BC,
although BS is 3 to 4 times
higher than BC undercurrent
U.K. aerosol conditions. In
1961, the UK established
a national air pollution
monitoring network,
called the National Survey,
monitoring black smoke and
sulphur dioxide at around
1,200 sites in the UK. (Bower
etal.,2009)
UK Trends
250
"200
£ 150
100
3 50
1955 1960 1965 1970 1975 1980 1985 1990 1995 2000
"•" Tg Annual BC Emissions ~*~ Annual Mean Concentration, ug/m3 (BS)
0.3
0.25 -3
0.2 «f
0.15 1
0.1 S
CD
0.05
Figure 5-7b. Comparison
of Ambient Black Smoke
Measurements (ug/m3,
annual average) with
Estimated BC emissions
(Tg) in the United Kingdom,
1955-2000. (Source: U.S. EPA)
724
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Observational Data for Black Carbon
m
E
4-
c
O "3
+= J"
03
8 2
O
U 1
U
DQ
0^
JAN88 JAN90 JAN92 JAN94 JAN96 JAN98 JANOO JAN02 JAN04 JAN06 JAN08 JAN10
Figure 5-8. Ambient BC Trends in Washington, D.C. The red data points and line represent daily and
annual average concentrations from Wednesdays (as a proxy for weekdays) and the black data points
and line represent daily and annual average concentrations from Saturdays (as a proxy for weekends).
The large dots represent the annual average concentration which is plotted at mid-year. This monitoring
site changed its sampling protocol from twice per week (Wednesday and Saturday) to once every 3 days
in August 2000. The apparent increase in BC concentration after January 2005 coincides with a carbon
analyzer upgrade resulting in total carbon measurements with a higher proportion of BC. (Source: U.S. EPA,
produced using data from VIEWS http://vista.cira.colostate.edu/views)
Note: The potential increase in reported EC measurements is described by White (2007), http://vista.cira.colostate.edu/
improve/Data/QA_QC/Advisory/da0016/da0016_TOR2005.pdf.
5.4 Trends in Ambient BC
Concentrations
5.4.1 Trends in Ambient BC Concentrations
in the United States and the United Kingdom
Measurement data necessary for assessing long-term
ambient trends in BC are limited even for the areas
with currently robust monitoring programs such as
the United States.13 However, although limited, some
information on changes in ambient concentrations
over longer time periods is available and these data
are useful in evaluating and corroborating emissions
trends. Since most BC is directly emitted rather
13 Assessment of longer term trends in BC is possible by analyzing
ice core and lake sediment data. These data reflect historical
archives from which BC concentrations can be estimated and used
to supplement more recently available direct ambient air quality
measurements. A discussion of these data and the corresponding
results is the focus of section 5.6 of this chapter.
than the formed chemically from precursors in the
atmosphere, ambient BC concentrations respond
directly to emissions changes. Figure 5-7a shows
the dramatic reduction in measured "Black Smoke"
(BS) in the UK since the 1950s. This dramatic decline
is attributable to a number of factors, including the
introduction of cleaner fuels and technologies, and
successful smoke control legislation (Bower et al.,
2009). Figure 5-7b overlays these BS measurements
and estimated BC emissions for the UK for the same
time period, revealing large estimated emissions
reductions corresponding to 80% of the reduction in
black smoke.
For more recent time periods, there is a great deal
more data available to assess ambient trends in
the United States than there is for longer-term
historical trends. A variety of measurements
from the IMPROVE and CSN networks, as well as
other monitoring locations, provide important
data for assessing recent changes in ambient BC
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Chapter 5
Elemental Carbon
Mar 1990-Feb 2004
1991 avg.=0.29 yg m '
Figure 5-9. Trends in BC at All IMPROVE Network Stations with Sufficient Data between 1 March 1990 and
29 February 2004. Marker size indicates the magnitude of the trend. Triangle direction and blue or red color
corresponds to the sign of the trend. Color saturation is proportional to the average concentration in 1991
with full saturation at twice the national median. The only urban site in Washington, D.C. is marked. Averages
in the bottom right corner exclude Washington, D.C. (Adapted from Murphy etal., 2011)
concentrations domestically. Figure 5-8 shows the
1988-2009 (22-year trend) for BC in Washington, D.C.
as measured by the IMPROVE program. IMPROVE's
urban Washington, D.C. monitoring site has one of
the longest BC monitoring records in the United
States. These data are presented as separate time
series due to a change in the frequency of sampling
and with a separation at the end of 2004 due of an
upgrade to newer analytical equipment. The trends
show a substantial two decade decline in ambient BC
concentrations. The percent change in Washington,
D.C.'s BC on Wednesdays and Saturdays was 62%
and 49% respectively, based on a comparison of
average levels in 1989-1991 compared to 2002-
2004, and 54% and 40% comparing 1989-1991
with 2007-2009. The higher BC concentrations and
more substantial BC reductions during the week as
compared to the weekend may correspond to the
influence of the reduction in diesel emissions.
Nationwide reductions in average BC concentrations
have also been observed in rural areas during this
same time period (Figure 5-9). BC concentrations
in the rural United States decreased by over 25%
between 1990 and 2004. Although not shown in
this figure, percentage decreases were much larger
in winter, suggesting that emissions controls have
been effective in reducing concentrations across
the entire United States (Murphy et al., 2011).
The large 22-year urban BC decline illustrated for
Washington, D.C. may in fact be larger than the
overall estimated nationwide reductions in BC and
direct PM2.5 emissions in the United States described
in Chapter 4, and is the result of area-specific
emissions dominated by certain emissions sectors.14
Figure 5-10 juxtaposes estimated annual average BC
concentrations in the San Francisco Bay with annual
consumption of diesel fuel in California (Kirchstetter
et al., 2008).15 Kirchstetter notes that the contrast in
the trends in BC concentration and diesel fuel use
is striking, especially beginning in the early 1990s
when BC concentrations began markedly decreasing
despite sharply rising diesel fuel consumption.
This contrast suggests that control technologies
to reduce BC emissions have been successful (see
14 As stated in Chapter 4, national BC emissions decreased by 79%,
30%, and 25% for on-road gasoline, on-road diesel, and nonroad
diesel sources, respectively, from 1990 to 2005. Also, Chapter 8
(Table 8-1) shows that 45% of the on-road BC reductions are due
to gasoline vehicles. Thus a combination of on-road gasoline and
diesel are each potential contributors to the lower BC in our nation's
capital and other urban areas.
15 BC was estimated using Coefficient of Haze (COM) measurements,
which are shown to be highly correlated with optical BC. See
Appendix 1 for further details regarding COHs.
726
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Observational Data for Black Carbon
O)
O
CD
s
ro
CO
o
I
LL
C
CD
w
00
U
DO
c
O
03
ro
X
3--
2 --
1 --
-- 60 =•
--50
--40
70
Q.
E
o
O
-- 30
--20
-- 10
cu
Q
.ro
'c
,g
"ro
1965
1975
1985
1995
2005
Figure 5-10. Estimated Annual Average Ambient
BC Concentrations in the San Francisco Bay Area vs.
Diesel Fuel Consumption. BC is shown as black dots.
California on-road (blue) and nonroad (gray) diesel fuel
consumption are shown as triangles and diamonds.
(Kirchstetteretal., 2008)
1.5
1.0
0.5
0.0
1999
2001
2003 2005 2007 2009 2011
Figure 5-11. Ambient BC Trends in Boston (Harvard
School of Public Health location). The blue squares are
monthly averages and red line is the annual average of BC
concentrations as directly reported by an Aethalometer.
(These data were not adjusted using correction
algorithms described in Appendix 1.) (Source: U.S. EPA,
data courtesy of Harvard School of Public Health)
Chapter 8). Similarly, Figure 5-11
shows a data set from Boston,
MA which displays a decline in BC
concentrations during the period
2000-2009. These data are presented
as monthly average concentrations
to help reveal the underlying trend,
but much higher temporal variability
exists in the 5-minute or even hourly
concentrations which respond to
patterns in nearby mobile emissions
and can even identify when a diesel
truck passes by. The decline of BC
concentration at this site has been
attributed to diesel retrofits in
Boston, but is no doubt also reflective
of fleet wide changes in emissions
especially due to diesel emissions
standards (U.S. EPA, 2004c).
Figure 5-12 shows that BC
concentrations have declined 32%,
on average, for a 15-site subset of
EPA's national urban CSN monitoring
sites with the longest historical
record. The figure shows the range of
monthly average BC concentrations
represented by CSN EC measurement
data. Because EPAtransitioned its
urban EC monitoring to the IMPROVE
protocol, CSN EC measurements
have been produced by two different
monitoring protocols. As discussed in
section 5.2, EC may vary by a factor of
two and in fact, that may also be true
for these two data records. Therefore,
the concentration data for these sites
(which were all part of the first group
which switched from the older CSN
NIOSH-type monitoring protocol) are
shown as reported after May 2007
while the earlier data are adjusted
to be IMPROVE-like.16 The details
about this adjustment are described
in Appendix 1. In general, the CSN
sites and trend are representative
of neighborhood, urban-wide, and
regional-scale emissions influences
and may not necessarily reflect
local scale emissions changes.
Nevertheless, the range of monthly
average concentrations among the
group of 15 urban locations illustrates
16 As discussed in Appendix 1, the adjustments
are based on parallel measurements at 14 urban
IMPROVE and 168 urban CSN locations between
2005 and 2011; the adjustment and trend prior
to 2005, therefore, is more uncertain.
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Chapter 5
CO
E
1? 2
d
g
U—'
ro
lr
-------
Observational Data for Black Carbon
+ AO
60
40
on
20
HI
c)
Zeppelin
84
89
94
99
Year
Figure 5-13. The Annual Mean BC Concentrations Measured at Alert (a). Barrow
(b), and Zeppelin (c) and Split into Contributions from the Four Transport
Clusters. The annual mean concentrations measured at Alert (a), Barrow (b), and
Zeppelin (c) are split into contributions from four transport clusters. The solid line
shows the linear trend through the measured concentrations. The circles show the
annual mean BC concentrations when the cluster-mean concentrations are held
constant over time (means over the first three years). This line is influenced only by
changes in the frequencies of the four clusters. The dashed line shows the linear
trend of these data. (Hirdman et al., 2010)
clusters (Arctic Ocean - AO, North America - NA,
Pacific-Asia - PA, and west northeast Eurasia - WNE)
have been stable or decreasing over the time periods
in this study.
5.5 Remote Sensing Observations
Measurements from satellite and ground-based
remote sensing are useful in describing global
aerosol and, in particular, BC absorption. Satellites
systems designed with aerosol remote sensing
capability include MODIS and MISR on Terra and
Aqua, as well as GLAS and CALIPSO lidars which
describe aerosol layer heights and other satellite
instruments such as the Total Ozone Mapping
Spectrometer (TOMS) (Winkler et al., 2007). The
ground-based remote-sensing Aerosol Robotic
Network (AERONET) has provided information
on aerosol distribution, seasonal variation and
absorption properties since 1963 (Holben et al., 1998;
Kahn et al., 2010; 2007; 2009; Kazadzis et al., 2009).
Unlike spatially discrete
ambient BC monitors, remote
sensing observations are
global and thereby offer
greater spatial surface
coverage of BC levels and
provide important estimates
of BC where surface ambient
measurements are not
available. While remote
sensing does not necessarily
characterize surface
concentrations, it provides
important information
on spatial variability of
concentrations in BC
and aerosols throughout
the total atmospheric
column. Combining these
new data sources with
traditional ground based
(ambient) measurements
has been used to derive
the complete aerosol effect
on the environment and
climate (Falkeetal., 2001;
Husar, 2011). Integrated
data sets of aerosol based
extinction have relied
heavily on AERONET sun
photometer measurements
in remote locations with low
concentration and relatively
homogeneous aerosol
(Kaufman et al., 2001), while
downwind of pollution or dust sources they have
relied on MODIS characterization of the aerosol
spatial distribution over the ocean and dark surfaces
(Remer et al., 2002) and on TOMS over bright
surfaces (Torres et al., 2002).
AERONET derived estimates of total column aerosol
optical depth (AOD) at 4 wavelengths (440, 670, 870
and 1020) can further characterize other aerosol
optical properties, including an estimate of Aerosol
Absorption Optical Depth (AAOD) throughout the
absorption spectrum (Holben et al., 1998; Dubovik
and King, 2000). Similarly, aerosol measurements
from the Ozone Monitoring Instrument (OMI) of
TOMS also provide a measure of AAOD.
Koch compares estimated AAOD for 1996-2006
based on AERONET with OMI satellite retrievals for
2005-2007 (Koch et al., 2009; Torres et al., 2007).
Koch notes that the two data sets broadly agree
with one another. However, the OMI estimate is
larger than the AERONET value for South America
(with UV sensitive biomass combustion) and
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Chapter 5
+ North America
+ Europe
+ Southeast Asia
" o South America
. o South Africa
+ Other
0.1 0.2 0.5 1 2
AERONET
5 10
20.0
6.0
5.0
4.0
3.0
2.0
1.0
0.5
0.2
0.1
0.0
Figure 5-14. Aerosol Absorption Optical Depth (AAOD) from AERONET (1996-2006) and OMI (2005-
2007). (a) Aerosol absorption optical depth, AAOD, (xlOO) from AERONET (at 550 nm), (b) OMI (at 500 nm)
and (c) scatter plot comparing OMI and AERONET at AERONET sites. (Koch et al., 2009)
440,670
440,870
440,1020
V-
-------
Observational Data for Black Carbon
(a)
Q
O
2.5 1
2.0
1.5
1.0
0.5
0.0
25
20
g
15 a
10 9
J FMAMJ JASOND
Month
(b)
2.5
2.0
1.5
1.0
0.5
0.0
J FMAMJ JASOND
Month
Figure 5-16. AERONET AOD and AAOD as a Percentage of AOD. (a) Goddard Space Flight Center (GSFC)
in eastern United States and (b) Beijing, China from May 2005 to April 2009. Red dots represent AOD
(shown on left vertical axis). Both AOD and AAOD are at 870nm. Horizontal axis shows the month of the
year. The purple and black lines are the monthly average AOD and percent absorbing aerosol. Note that at
GSFC, AOD is highest in the summer (associated with secondarily formed aerosols and winter-time AAOD
observations are not available for GFSC because the calculation is not possible at low AOD values. (Source:
AERONET data are based on Version 2, Level 2.0 inversion products with permission of Brent Holben, NASA
and Hong-Bin Chen Chinese Academy of Sciences)
25
20
15 O
10 §
smaller for Europe and Southeast Asia, which are
dominated by BC. The AERONET AAOD and OMI
observations qualitatively agree with the ground
level concentrations of BC for the United States,
Europe, and Asia presented in Figure 5-14, and
clearly increase the spatial characterization of
aerosol absorption. As discussed below, aerosol
absorption may not necessarily be associated with
anthropogenic source emissions.
Multi-wavelength instruments, such as AERONET,
can also characterize the wavelength dependence of
absorption (often expressed as Absorption Angstrom
Exponent, or AAE) to provide an indicator of the
absorbing aerosol mixture. Using pairs of wavelength
specific absorption measurements, Russell et al.
(2010) find AAE values near 1 (the theoretical value
for BC) for AERONET-measured aerosol columns
dominated by urban-industrial aerosol, larger AAE
values for biomass burning aerosols, and the largest
AAE values for Sahara dust aerosols. Using these
observations from multi-wavelength sensors can
help distinguish the types of absorbing aerosols
(Figure 5-14). It also demonstrates that the global
AAOD observations presented in Figure 5-14 do
not exclusively represent BC from anthropogenic
sources.
17
17 The illustrative remote sensing observations presented in section
5.5 will be considerably strengthened when geostationary GLORY
satellite with broad spectrum solar sensors to determine the global
distribution of aerosol and cloud properties is deployed. Glory will
provide 9-wavelength single-scattering albedo (SSA), AOD, AAOD,
and AAE, as well as shape and other aerosol properties (Mishchenko
et al., 2007; Russell et al., 2010).
While a common limitation of remote sensing
(which depends on solar light) is its general
representativeness of daytime and cloudless sky
conditions, AAOD is additionally only representative
of higher extinction periods required to make the
needed absorption calculations. Consequently,
AAOD for the United States and Europe is not
based on measurements during the winter when
atmospheric extinction is lower than the minimum
computational threshold. Similarly, AAOD are
not as well represented during the monsoon
periods in Asia when AOD measurements are not
available. These issues may be partially addressed
by using seasonally or monthly weighted averages.
Figure 5-15 illustrates the issue of incomplete data
records and the contrast of AAOD levels across the
globe.
Figure 5-16 presents AOD and AAOD as a
percentage of AOD by date for two example
locations from AERONET (Goddard Space Flight
Center, GSFC in MD, and Beijing, China). The fraction
of AOD that is estimated to be absorbing is lower
at GSFC, but those data are not available during the
winter months at GSFC due to insufficient AOD to
calculate the absorbing portion. This is typical of the
eastern United States and other U.S. locations where
sufficient AOD only exists in the summer and which
principally results from secondarily formed mostly
scattering aerosols. Average percent absorbing
aerosol for GSFC derived from AERONET is about
6-7%, but this does not represent winter-time
conditions when BC may have its highest values.
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Chapter 5
In contrast, AOD is sufficiently high year-round in
Beijing and is several times the amount detected
at GSFC; the Beijing absorbing portion (black line)
ranges from about 8-9% in the summer to 15% in
the winter. For the 7 months of the year where both
locations provide data, the ratio of monthly average
AAOD values between Beijing and GSFC ranges from
2 to 7. However, without year-round AAOD data for
GSFC, the estimate of its annual average absorbing
aerosol and its annual average ratio to Beijing is very
uncertain. Thus, the lack of wintertime absorption
measurements limits the value of remote sensing
estimates to expand the spatial extent of ground
level measurements for model evaluation and
corroboration of emissions inventories.
5.6 BC Observations from Surface
Snow, Ice Cores, and Sediments
Snow and ice cover approximately 7.5-15% of the
Earth's surface, depending upon the time of year
(Kukla and Kukla, 1974). The sunlight that reaches
the snow surface typically penetrates about 10-20
cm into the snow, with the topmost 5 cm receiving
the most sunlight and where light-absorbing
impurities can significantly alter the amount of solar
energy reflected by the snowpack (e.g., Galbavy
et al., 2007). Black carbon measurements in snow,
and related surface reflectivity measurements, are
critical to accurately estimate climate forcing due to
snow-bound BC. In addition, ice core measurements
of BC provide an important record of natural and
anthropogenic BC emissions transported to snow-
covered regions. Lake and marine sediments also
pose an opportunity to derive historical trends in
BC emissions prior to the point of time when air
monitoring data are available.
5.6.1 Measurement Approach
Measurement of BC in snow or ice is a laborious
process that begins with careful manual collection of
snow or drilling an ice core. A sample of snow or ice
is then melted and BC is quantified through several
analytical approaches. The majority of researchers
filter the melted snow or ice, collecting BC to the
filter matrix and estimating BC by observing how
light at certain wavelengths is absorbed by the
particles (Grenfell et al., 1981; Clarke and Noone,
1985) or through a thermal or thermal-oxidative
process (Ogren et al., 1983; Chylek et al., 1987).
In addition, one newer approach avoids filtering
the snow and quantifies BC by laser-induced
incandescence (McConnell et al., 2007). The mass
of the sample meltwater is measured and the final
concentration units are usually in mass of BC per
mass of snow or ice (e.g., ng BC/g snow).
Quantification of BC in sediments is an emerging
field of study. The measurement technique is
more complex than for snow or ice samples, as BC
particles are embedded in sediment material that
contains significant amounts of organic material.
The sampling process usually involves extracting
a sediment core and then slicing the core into
layers. The BC particles are subsequently isolated
for a given sample by applying a series of chemical
and/or thermal treatments designed to remove
non-BC material (Lim and Cachier, 1996; Kahn
et al., 2009; Smith et al., 1973). Once the non-BC
material is removed to the degree possible, BC
concentrations are quantified via similar techniques
utilized in ice core or ambient samples - measured
by light absorption or through thermal processes.
Microscopic analysis of carbon particles has also
been employed to qualitatively determine the
source type from the particle shape and surface
texture (Kralovec et al., 2002; Smith et al., 1973).
5.6.2 Surface Snow Data
Measurements of BC in the shallow surface layer
of snow have been conducted since the 1980s
by research teams at locations throughout the
Northern Hemisphere and in Antarctica, although
the measurements were sporadic (Figure 5-17).
Two large field studies, Clarke and Noone (1985)
and Doherty et al. (2010), significantly boosted the
number of sampling locations during two windows
of time (1983-1984, 2006-2009). However, even
the highest number of measurement locations (55
sites in 2009) provides sparse geographic coverage
of data, considering the high degree of spatial
variability in BC concentrations. Recent model
estimates by Flanner et al. (2007), seek to fill in
the missing measurement gap with predictions of
surface snow BC concentrations in the northern
hemisphere, estimating values ranging five orders of
magnitude (<1 to >1000 ng BC/g snow).
Recent surface snow results from Doherty et al.,
(2010) show that BC concentrations range over
an order of magnitude in remote areas of the
Northern Hemisphere (Figure 5-18). Even higher
BC values in snow were reported for the Tibetan
Plateau and throughout western China, up nearly
another order of magnitude (Ming et al., 2009;
Xu et al., 2006). BC removal from the atmosphere
is primarily driven by precipitation (Ogren et al.,
1984), thus BC concentrations in snow or ice are a
function of the atmospheric concentration of BC
above the surface and the frequency and amount of
snowfall in a particular area. For example, Xu et al.
(2009b) noted that BC concentrations on the Tibetan
plateau were high during nonmonsoon periods with
low precipitation, which they related to regional
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Observational Data for Black Carbon
60
50
o
fi
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40
Antarctica
I Asia
Northern Europe
North America
i Arctic Ocean
I Russia
i Greenland
oooooooocoooooooooooaiaicnaicncncnaicncnoooooooooo
cncriCTiCTjcjicTjcncnaicjicjicncnaia'iCTicrjCTicna'ioooooooooo
Figure 5-17. Locations of BC Measurements in Surface Snow and Shallow Snow Pits (snow pits are
indicated for each year covered in the pit depth). (Sources: U.S. EPA, based on data reported in Cachier and
Pertuisot (1994), Cachier (1997), Chylek et al. (1999; 1987), Clarke and Noone (1985), Doherty et al. (2010),
Grenfell et al. (2002; 1981; 1994), Hagler et al. (2007a; 2007b), Hegg et al. (2009; 2010), Masclet (2000), Ming
et al. (2009), Perovich et al. (2009), Slater et al. (2002), Warren and Clarke (1990), Warren et al. (2006), and Xu
etal. (2006))
Greenland Arctic Arctic Svalbard Subarctic Northern Arctic
Ocean Canada Canada Norway Russia
Siberia
Figure 5-18. BC Concentrations in Surface Snow in Arctic and Subarctic Areas of the Northern
Hemisphere. (Source: Derived from recent measurements reported in Doherty et al., 2010)
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Chapter 5
particulate pollution ("Asian Brown Cloud") elevating
during the dry nonmonsoon period and then highly
concentrating the infrequent precipitation with
impurities. An additional important factor, discussed
by several studies (Doherty et al., 2010; Planner et
al., 2007; Grenfell et al., 2002; Xu et al., 2006) is the
potential increase in surface snow BC levels when
melting snow leaves behind BC particles, further
darkening the topmost layer of snow.
It is important to note that certain non-BC particulate
species have been shown to absorb light when
deposited to snow or ice. While dust is not as strong
of a light absorber per unit mass as BC, dust can play
a significant role in reducing snowpack reflectivity
at high concentrations (Warren and Wiscombe,
1980). In addition, BrC in snow has been suggested
to significantly absorb light (Doherty et al., 2010).
Given that studies suggest that organic material in
snow may undergo chemical transformation and loss
from the snowpack due to sunlight-driven reactions
(Hagler et al., 2007a; Grannas et al., 2004), BrC may
absorb light to an even greater degree in fresh
precipitation than what has been measured in aged
snow samples. However, neither Grannas et al. nor
Hagler et al. specifically measured BrC or the time
evolution of light absorption.
5.6.3 Ice Core Data
Measurements of BC in ice cores are critical to
understanding the longer-term trends of human
influence on snow reflectivity. Ice cores, produced by
drilling into permanent ice and carefully extracting
a column of ice, have been collected and analyzed
for BC at a number of locations in the Northern
Hemisphere (Figure 5-19). In addition, an Antarctic
BC ice core record spanning the past two and a
half thousand years has just been completed as
part the National Science Foundation WAIS Divide
deep ice core project (Ross Edwards, personal
communication). The ice cores with continuous BC
data available primarily cover the past few hundred
years, with the exception of the Dye 3 ice core in
Greenland and the WAIS Divide core in Antarctica
which extend back several thousand years. The
layers of the ice core are dated using several
strategies, including measuring certain chemical
species with known seasonal variation, looking for
certain known historical events that had unique
Greenland
Ice Sheet
Alps
Himalayas and
Tibetan Plateau
1
GISP2
| Camp Century
1
Antarctica
>6000
EUROCOREl
Colle Gnifetti
Col du Dome |
Mt. Muztagh Ata
Tangula
East Rongbuk, Mt. Everest
Noijin Kangsang
Zuoqiupu
PLZ4
ERIC2002C , East Rongbuk, Mt. Everest
East Rongbuk, Mt. Everest
WAIS Divide
4500 3000
Years before present
1500
Figure 5-19. BC Ice Core Records Worldwide Labeled by Their Identifying Name. The extent of the bars
(light blue and/or black) shows the time covered by the depth of the ice cores, with the black regions
representing sections of the ice core that had BC concentrations reported. (Source: U.S. EPA, based on
data reported in Cachierand Pertuisot (1994),Chyleketal. (1987; 1992; 1995), Ross Edwards (personal
communication regarding WAIS Divide ice core), Kaspari etal. (2011), Lavanchy etal. (1999), Leg rand et
al. (2007), Liu et al. (2008b), McConnell et al. (2007), McConnell and Edwards (2008), Ming et al. (2008),
Thevenon et al. (2009), and Xu et al. (2009a; 2009b))
734
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Observational Data for Black Carbon
chemical signatures (e.g., volcanic eruptions, nuclear
explosions), and observing the visible layering of ice
throughout the core (e.g., Hammer, 1978).
The concentrations of BC in a certain ice core reflect
the past atmospheric concentrations above the
region, which in turn relate to short- and long-
distance transport of BC emissions. Thus, the ice core
results vary from location to location. For example,
on the remote Greenland Ice Sheet, McConnell
(2010) showed a peak in BC concentrations in the
1910-1920 time range, decreasing in concentration
from that point to present day. Meanwhile, ice
cores in the European Alps show BC concentrations
increasing significantly past the 1910-1920 period,
with highest concentrations recorded in the 1950-
1960 time frame (Lavanchy et al., 1999; Legrand et
al., 2007). Finally, Xu et al. (2009b) and Ming et al.
(2008) reveal variable results for multiple shallow
ice cores collected in the Himalayas and Tibetan
Plateau that date from the 1950s to 2004: several ice
cores have highest BC levels in the 1960s and lower
levels from that point forward, while another ice
core had continuously increasing levels until present
day. Studies of ice core data collected to date find
associations between elevated BC and human
activities; however, the trends vary significantly by
location.
5.6.4 Sediment Data
With ice core records only available in remote,
high-altitude locations in the world, undisturbed
lake sediments provide additional spatial coverage
of BC historical trends and may demonstrate higher
associations with local emissions. In addition,
deep ocean marine sediments reveal ancient BC
trends related to natural emissions. Similar to ice
cores, BC records in sediments initiate from the
deposition of BC from the atmosphere, which relates
to the atmospheric transport of BC emissions to a
particular location. After depositing to the surface of
a water body, the BC particles eventually transport
downwards and, if the sediment is undisturbed, may
form a permanent archive in the layers of sediment.
Lake sediment BC records have been quantified for
several interior lakes in North America, including
Lake Michigan, dating 1827-1978 (Griffin and
Goldberg, 1983), and four lakes located in the
Adirondacks of New York, dating 1835-2005 (Husain
et al., 2008). Total carbon particles, associated with
specific sources by particle shape, have also been
measured in Lake Erie sediments, dating 1850-1998
(Kralovec et al., 2002). Historical BC records have
also been obtained for a number of lakes in the Alps
of northern Slovenia (Muri et al., 2002; 2003) and
in ancient marine sediments, aged approximately
100 million to 5000 years before present, spanning
southern to far northern latitudes of the Pacific
Ocean and at several locations in the Atlantic Ocean
(Smith et al., 1973).
The findings by Smith et al. (1973) reveal an
approximate 10-fold increase in ancient BC
deposited levels moving from the equator
northward to 60°N (bisecting Canada), which they
related to the increase in natural wildfire emissions
moving from the equator northward. These trends
lay the base pattern of deposited BC, to which
anthropogenic emissions of BC would be added.
Focusing on sediment findings that closely relate
to U.S. emissions, Figure 5-20 presents estimates
of atmospheric BC derived from sediment core
measurements in the Adirondack region of New
York State for deposition from approximately 1835
to 2005 (Husain et al., 2008) and overlays these
estimates with long-term U.S. BC emissions data
developed by Novakov (2003). The derived BC
ambient estimates are well correlated with the
historical BC emissions estimates for fossil fuel
combustion in the United States, and Husain et
al. (2008) attributed the decrease from 1920-2000
to reduction in BC emissions from U.S. fossil fuel
combustion.
The ambient BC determined from Adirondack lake
sediments by Husain et al. (2008), shown above,
can be compared with records obtained from Lake
Michigan sediments from Griffin and Goldberg
- -A- -BC emissions
4!
(i
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1840 1880 1920
Year
1960
2000
Figure 5-20. Atmospheric BC determined by Husain et
al. (2008), for the Adirondack Region from 1835 to 2005.
The measurements are compared with U.S. BC emissions.
(Novakov etal., 2003)
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Chapter 5
(1983) and from Lake Erie sediments (Kralovec et al.,
2002). While the multiple archives in multiple North
American lake sediments reveal a significant increase
in deposited BC levels after the late 1800s, the Lake
Michigan time series shows an apparent peak in the
record around 1940-1960, while the Adirondacks
show a peak near 1910-1930, and the Lake Erie
record shows a peak in 1970-1980. It is uncertain
why the multiple sediment records in North America
contrast in the timing of the peak BC concentration;
possible explanations may include differences in
local versus regional source contributions, dating
methodology, carbon measurement approaches, and
sediment deposition processes.
ca
:
1800
2000
Figure 5-21. Annual Average Concentrations of (a) BC and VA and
(b) BC and Non-Sea-Salt Sulfur (nss-S). The gray shaded region
(between the black and blue dotted line) in the top figure represents
the portion of BC attributed to industrial emissions, not boreal forest
fires. (Source: Adapted from McConnell et al. (2007))
5.6.5 Arctic BC Snow and Ice Data - Source
Identification
Impacts of BC emissions on the Arctic are of
particular interest given the climate-sensitive nature
of the region. BC emissions from particular source
types or regions and transport to the Arctic have
been explored through modeling studies and field
measurements. This section discusses the findings
in observational BC data from Arctic snow and ice.
Connections between snow and ice BC data and
source types are generally made by measuring
additional species in the snow (i.e., ions, metals,
organics, and isotopes) and comparing trends
between the multiple data sets.
Historical trends in Arctic ice cores collected on the
Greenland Ice Sheet improve our understanding of
the historical impact of anthropogenic
and natural emissions of BC on the Arctic.
McConnell and Edwards (2008) and
McConnell et al. (2007) provide monthly-
resolution BC data in ice cores on the
Greenland Ice Sheet. Similar to the lake
sediment findings for the Adirondack
Mountain region, the maximum BC
concentrations in Arctic ice in the past
hundred years occurred in the early
1900s corresponding to increases in
a number of species associated with
industrial emissions (e.g., cadmium,
cesium, thallium, lead). McConnell et al.
(2007) compare vanillic acid (VA), non-
sea-salt sulfur (nss-S), and BC trends to
apportion the BC due to industrial versus
forest fire emissions (see Figure 5-21). VA
is considered an indicator of forest fire
emissions, while nss-S relates to industrial
emissions and volcanic eruptions. In the
postindustrial era, BC anthropogenic
emissions contributed roughly 50-80%
of the total BC loading in the ice during
early 1900s and over past few decades
the industrial input was on the order of
20-50% (estimated from Figure 5-21,
originally published in McConnell et al.,
2007). While nss-S correlated highly with
the increasing BC during the late 1800s
to mid-1900s, the trends did not match
later; this may be related to changes in
industrial emission factors. This study
associates the high BC concentrations
in the early 1900s with North American
fossil fuel emissions and suggests that
Asian emissions may play an important
role past the mid-1900s.
-
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Observational Data for Black Carbon
Two recent studies attempted to attribute BC
in snow to specific sources by collecting a large
number of surface snow samples throughout the
Arctic, which they measured for detailed chemical
composition (Hegg et al., 2009; 2010). In Hegg
et al. (2010), statistical analyses revealed that the
measured species could be grouped into four unique
factors with source-defining chemical characteristics
(for example, sodium and chloride indicating a
marine environment), which the authors labeled as
marine, boreal biomass, crop and grass biomass,
1.0
0.0
Boreal
biomass
a
150
(b)
—
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\-
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S T — ^ _
V %, % V V>
Sampling location
Figure 5-22. Sources of BC in Arctic Snow, (a) Fractional
source contributions to Light Absorbing Aerosol (LAA) snow
concentrations in Siberia (Pevek, Billbino, Cherskiy, Tiksi, Yakutsk),
the Greenland Ice Sheet, and the North Pole, (b) The box and
stem plots represent concentrations of LAA at each location, with
error bars indicating the 95% confidence interval. LAA values
are derived from a light absorption technique which converts
to a mass estimate using calibration factors and are generally
equivalent to BC values, although non-BC absorbing aerosols
could bias the estimate high. (Source: Hegg et al., 2010)
and pollution. It should be emphasized that the
"marine" category represents air masses with
an ocean-like chemical signature (i.e., sea salt),
which may also include emissions from other
sources (biomass or fossil fuel combustion) that
transported over the ocean and mixed with sea
spray. Depending on the location of the sample
within the Arctic and time of year, the estimated
contribution from these four sources varied
considerably (Figure 5-22). In Siberia, emissions
from biomass burning were significant drivers of
BC and other absorbing species. However,
on the Greenland Ice Sheet and at the
North Pole, pollution and crop/grass
biomass were found to be the primary
sources.
5.7 Limitations and Gaps in
Current Ambient Data and
Monitoring Networks
The primary limitation in existing
ambient monitoring data is the sparse
geographic coverage of existing BC
monitoring locations. There are parts of
the world where there currently are no
measurements; and where they do exist,
the measurements are not archived into
a consolidated database. The differences
in average BC concentrations between
countries (global scale), among regions
(regional scale) and also within cities
(local scale) are all much larger than the
differences across monitoring methods.
These geographic variations are also larger
than the inter-annual changes that may
occur within a 5-to-10-year period. To
help develop and corroborate emissions
inventories and to evaluate global models
(see Chapters 4 and 2, respectively),
additional ambient measurements are
needed at more locations. Existing
geographically dense filters used to
produce PM2.5 measurements in the United
States (and elsewhere if available) can be
used to cost-effectively supplement the
measurements from more specific and
expensive BC monitors. Also currently
there are insufficient measurements
characterizing the BrC component of OC.
The addition of more multiple wavelength
optical analyzers or use of optical
measurements from existing PM2.5 filter
samples would be useful (Hecobian et al.,
2010; Chow etal.,2010b).
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Chapter 6
Benefits of Reducing Black Carbon
Emissions
6.1 Summary of Key Messages
• Mitigation of BC offers a clear opportunity:
continued reductions in BC emissions can provide
significant near-term benefits for climate, public
health, and the environment.
• It is generally impossible to reduce BC in isolation
from other co-pollutants: most BC mitigation
strategies involve reductions in total emissions of
fine particles, including other PM2.5 constituents.
Fortunately, all control measures that reduce PM2.5
pollution—including BC and other constituents—
will achieve substantial health benefits. Therefore,
mitigation strategies targeting BC that also reduce
total direct PM2.5 emissions could potentially result
in hundreds of thousands of avoided premature
deaths each year. The health benefits alone may
be large enough to justify mitigation in many
regions and sectors.
- Because a variety of PM constituents are
associated with adverse health impacts,
significant health benefits are anticipated
to result from reducing exposure to PM
containing BC, regardless of the precise
chemical composition of the emissions
mixture.
- Controls on direct PM2.5 emissions (including
BC) can be particularly beneficial since they
are more co-located with population than
other PM components. According to U.S. EPA
estimates, the benefits from control of direct
PM25 emissions are 7 to 300 times greater
than the benefits per ton estimated for
reductions of PM precursors such as NOX and
SOX.
- The magnitude of human health benefits of
emissions reductions depends both on how
much exposure is reduced and the size of
the affected population. The largest health
benefits from PM2 5 including BC control
strategies will be achieved in areas near the
emissions source and where exposure affects
a large population.
- Programs to reduce PM25 in the United States
and in other developed countries have greatly
reduced the negative health impacts of PM2.5,
including BC. New programs introduced for
mobile and stationary sources will continue
to reduce PM25-related health impacts in
the United States and the countries that
implement them over the next several
decades.
- The largest opportunities for achieving health
benefits of BC mitigation measures are in
lesser developed countries, due to high
emissions co-located with large populations,
particularly in South and East Asia.
Estimating the climate benefits of mitigation
strategies is more challenging than health
benefits estimation because key scientific issues
remain unresolved. Despite uncertainties in the
magnitude of effects, the literature suggests
climate benefits could be achieved through some
mitigation measures.
- BC reductions are expected to reduce
the rate of warming soon after measures
are implemented, though the full climate
response may take several decades to fully
manifest. In sensitive regions (e.g., the
Arctic), or in regions of large emissions (e.g.,
South and East Asia), additional benefits
may include slowing the rate of ice, snow,
and glacier melt, and reversal of adverse
precipitation changes.
- Achieving climate benefits is most likely to
result from the control of emissions from
BC-rich sources such as diesel engines,
cookstoves, brick kilns and coke ovens.
- BC mitigation alone cannot change the long-
term trajectory of global warming, which
is driven by CO2 emissions. However, as
illustrated by a recent UNEP/WMO analysis,
controlling current BC and long-lived GHGs
in concert would greatly improve the chances
of limiting global temperature rise below 2°C
relative to pre-industrial levels.
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Chapter 6
- BC mitigation may have particular benefits
for the Arctic and the Himalayas, two regions
that are particularly sensitive to BC deposition.
In the Arctic, both global and high-latitude
BC reductions can provide climate benefits.
The most effective emissions reductions per
unit of emissions will come from within- or
near-Arctic sources, but overall reductions in
global BC emissions will be critical to slowing
the rate at which the Arctic is warming and
melting. Slowing the rate of change may
avoid reaching certain global tipping points as
discussed in earlier chapters. In the Himalayan
region, a range of local measures is likely to
provide climate benefits, including sizable
benefits for radiative forcing, ice/snow melt,
precipitation and surface dimming.
Limiting BC emissions also is expected to provide
a number of environmental benefits, including
improved visibility, reduced surface dimming,
reduced impacts on ecosystems, and less damage
to building materials.
More research is needed on the benefits of
individual control measures in specific locations
to support policy decisions made at the
national level. Research is also needed to design
approaches to valuing the climate impacts of BC
directly, and to incorporate those approaches
into useful metrics for evaluating policy decisions,
similar to the social cost of carbon (SCC).
6.2 Introduction
This chapter summarizes available information
regarding the potential public health, climate and
environmental benefits of reducing BC emissions,
both in general and from particular economic
sectors. The literature on the health benefits of
reductions in fine particles (including BC) is well-
developed, particularly in the United States where
emissions control programs have been evaluated
extensively. These analyses provide a high degree of
confidence that BC mitigation strategies will produce
significant public health benefits, not just from the
BC reduction, but also from the reduction in co-
emitted gaseous and particulate pollutants as well.
As BC effects on climate are less certain than on
health, there is also less certainty about the climate
benefits of BC mitigation. Although the body of
literature on climate benefits is limited, it provides
important insights regarding which strategies are
most likely to provide climate benefits. Non-climate
environmental benefits of BC reductions are also
less certain than health benefits, but the literature
on PM2.5 impacts provides important qualitative
information regarding the likelihood of such
benefits. It is important to note that a quantitative
assessment of the benefits of specific mitigation
measures was not conducted for this report due
to time and resource constraints. Further analysis
would be necessary to quantify the public health,
climate and environmental benefits of specific
BC reduction strategies, either individually or in
combination.
6.3 Public Health Benefits of
Reducing Black Carbon Emissions
All control measures that reduce PM2.5 emissions,
of which BC is a component, are virtually certain to
achieve health benefits. The adverse health impacts
of PM2.5 are well documented in the scientific
literature, previously discussed in Chapter 3 and
in EPAs recent PM ISA (U.S. EPA, 2009a). The well
established linkage between PM2.5 and health effects
is important, because mitigation strategies aimed
at reducing BC almost always involve reductions in
co-pollutants, including other PM constituents or
precursors, as well as BC. BC is emitted in a mixture
of other pollutants that are also associated with
negative health effects, including primary particulate
matter (including OC), precursors of secondary
particulate matter (SO2, NOX, NH3), and ozone
precursors (NOX, CO, VOCs). Reductions in all of
these pollutants can provide human health benefits.
Though the literature on differential toxicity of PM2.5
components and mixtures is currently inconclusive,
studies continue to provide evidence that many
PM2.5 constituents are associated with adverse health
impacts, as discussed in Chapter 3. Because of the
well documented health impacts of PM2.5, there is
high likelihood that BC mitigation strategies that
reduce PM25 will produce public health benefits. This
section summarizes what is known about the health
benefits of BC emissions reductions in the United
States and globally.
6.3.1 Health Benefits in the United States
Historically, the United States has been quite
successful in achieving significant PM2.5 reductions
through air quality protection programs such as
attaining the PM25 NAAQS and implementing
a variety of mobile source rulemakings. These
reductions yield large human health benefits. As
discussed in Chapter 4, however, emissions of
BC in the United States are still fairly substantial.
Additional control programs, largely those affecting
mobile diesel engines, are expected to achieve
further BC reductions by 2030 with significant public
health benefits (see Chapter 8). EPAs analyses of
these benefits provide useful information regarding
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the potential public health improvements that can
be achieved via strategies targeting direct emissions
of PM2.5, including BC. Below, we describe how
the benefits of policies that reduce atmospheric
pollutant loads are developed, and discuss in more
detail specific estimates of the health benefits that
PM2.5 mitigation policies can help achieve.
6.3.1.1 Methods for Estimating the Health Benefits
of Reducing PM2 5
Peer-reviewed methods for estimating the public
health benefits of PM2.5 emissions reductions
have been used in a variety of settings and the
sophisticated models and techniques developed for
application in the United States have been refined
by EPA over the course of several decades. These
methods rely upon atmospheric models to translate
emissions changes into concentration changes, and
epidemiologically derived concentration-response
functions to calculate the change in a health
endpoint attributable to the concentration change.
Valuation techniques are then used to quantify the
economic impact of the health benefits, with the
total monetized benefit calculated as the sum of
the values for all non-overlapping health endpoints.
Health benefits are often monetized using the
Value of a Statistical Life (VSL), which is determined
by studies of individuals' willingness to pay (WTP)
for reducing their risk of mortality. This approach
is the standard method for assessing benefits of
environmental quality programs and has been used
widely in EPA regulatory documents (e.g., U.S. EPA,
2006c), as well as in the peer-reviewed literature (e.g.,
Levy et al., 2009; Hubbell et al., 2009; Tagaris et al.,
2009).
Air pollution affects a variety of health endpoints, as
discussed in Chapter 3. For some, currently available
data are insufficient to enable quantification or
monetization of effects. Table 6-1 summarizes the
health endpoints that have been included in recent
EPA PM2.5 benefit assessments. The table indicates
which effects have been quantified and monetized
(left column), and which are discussed only
qualitatively (right column).
Overall, the PM2.5 control program in the United
States has proven highly protective of public
health. Multiple studies estimate significant
reductions in PM2.5-related mortality and morbidity
since the implementation of the Clean Air Act
(CAA). In a recent study, EPA also estimated that
programs implemented as a result of the 1990
CAA Amendments have avoided about 160,000
annual premature adult deaths by 2010, and will
avoid 230,000 by 2020 (Table 6-2). Direct PM2.5
reductions specifically have been associated
with 22,000 to 60,000 annual avoided premature
deaths between 2000 and 2007 (Fann and Risley,
2011). Improvements in particulate air pollution,
particularly in urban areas, have also been estimated
to contribute 15% of a 2.72 year increase in average
life expectancy among 211 counties between 1980
and 2000 (Pope et al., 2009).
6.3.1.2 The Potential Benefits of Further Mitigation
of PM2 5 (Including BC) Emissions
Despite significant improvements, the health burden
of PM2.5 in the United States is still substantial.
Based on 2005 air quality data and population,
Fann et al. (2011) estimated that about 130,000
annual premature deaths and 19% of all ischemic
Table 6-1. PM2.5 Health Endpoints Included in EPA's Regulatory Impact Analyses. (Source: U.S. EPA, 2006,
Table 5-2)
_ ... . ... .. . _.., .. ,,..-. ... Un-Quantified and Non-Monetized
Quantified and Monetized PM25 Health Endpoints pM Hea|th End ints
Premature mortality based on cohort study estimates and expert
elicitation estimates
Hospital admissions: respiratory and cardiovascular
Emergency room visits for asthma
Nonfatal heart attacks (myocardial infarctions)
Lower and upper respiratory illness
Minor restricted activity days
Work loss days
Asthma exacerbations (among asthmatic populations)
Respiratory symptoms (among asthmatic populations)
Infant mortality
Low birth weight, pre-term birth and other reproductive
outcomes
Pulmonaryfunction
Chronic respiratory diseases other than chronic bronchitis
Non-asthma respiratory emergency room visits
UVb exposure (+/-)
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Chapter 6
heart disease-related deaths nationwide were
attributable to PM2.5 exposure. Fortunately, a number
of additional PM2.5 control programs that will help
to control PM2.5 (including BC) over the next several
decades have already been adopted and are
expected to yield significant health benefits. In the
illustrative regulatory impact analysis conducted by
EPA in 2006 for the revised PM2 5 NAAQS, benefits
Table 6-2. Changes in Key Health Effects Outcomes in the United States Associated with PM2.5 Resulting
from the 1990 CAA Amendments. (Source: U.S. EPA (201 la) Table 5-6, adjusted from 2006$ to 2010$)
Year 2010
Year 2020
Health Effect Reductions
Incidence Avoided
Valuation
(millions 2010$)
Incidence
Avoided
Valuation
(millions 2010$)
Adult Mortality
Infant Mortality
Chronic Bronchitis
Acute Bronchitis
Acute Myocardial Infarction
Hospital Admissions, Cardiovascular
Lower Respiratory Symptoms
Upper Respiratory Symptoms
Asthma Exacerbation
Lost Work Days
160,000
230
54,000
130,000
130,000
45,000
1,700,000
1,400,000
1,700,000
13,000,000
$1,300,000
$2,100
$26,000
$66
$15,000
$1,400
$32
$45
$97
$2,200
230,000
280
75,000
180,000
200,000
69,000
2,300,000
2,000,000
2,400,000
17,000,000
$1,800,000
$2,700
$39,000
$100
$23,000
$2,200
$45
$65
$140
$2,900
Notes:
1. EPA's estimates of reductions in respiratory hospital admissions, respiratory emergency room visit, and minor restricted
activity days are not included here, because the 2011 report presented these estimates only as combined totals for both
PM2.5 and ozone.
2. Estimates reflect annual incidence avoided and valuation for reductions in direct PM2.5 and PM2.5 precursors.
Table 6-3. List of Benefits, Costs, and Benefit to Cost Ratios for U.S. Rules with Direct PM Reductions
(Billions 2010$).a (Source: U.S. EPA)
Rule (by Sector)
Annual Benefits'1 Annual Costs
Benefit/Cost
Benefit Year
Transportation
Light Duty Tier 2
Heavy Duty 2007
Non-road Diesel Tier 4
Locomotive & Marine Diesel
Ocean Vessel Strategy
Stationary Sources
2006 PM NAAQSC
Cement NESHAP
Stationary Spark Ignition RICE NESHAP
Stationary Compression Ignition Engine NESHAP
$34
$92
$105
$9.9-23.8
$119-292
$22
$8.3-20.1
$0.52-1.2
$0.95-2.3
$7.2
$5.5
$2.5
$0.8
$3.4
$7.1
$1.01-1.04
$0.26
$0.37
4.7
16.7
41.5
13.1-31.4
35.5-87.1
3.1
8.2-19.4
2-4.7
2.5-6.2
2030
2030
2030
2030
2030
2020
2013
2013
2013
a Rules include a combination of direct PM2.5 and PM precursor reductions. These estimates have been adjusted from the
dollar years in the original analysis to 2010$.
b 3% discount rate used for benefit estimates.
c Estimates of benefits and costs for the PM NAAQS are illustrative since individual states will make the decisions about
actual control strategies implemented to comply with the NAAQS.
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Table 6-4. Direct PM2.5 National Average Benefits per Ton Estimates by Source Category for the United
States (3% Discount Rate, Thousands of 2010$). (Source: http://www.epa.gov/air/benmap/bpt.html, based
on method described in Fann et al., 2009)
,. _ Monetized Benefit Monetized Benefit Monetized Benefit
bource category Per Ton in 201 5 Per Ton in 2020 Per Ton in 2030
Area Source
Pope etal. (2002)
Laden etal. (2006)
$360
$880
$400
$970
$480
$1,200
Mobile Source
Pope etal. (2002)
Laden etal. (2006)
$280
$680
$300
$740
$360
$890
EGUandNon-EGU
Pope etal. (2002)
Laden etal. (2006)
$230
$560
$250
$610
$290
$710
Notes:
1. These estimates have been adjusted from 2006$ to 2010$.
2. These are U.S. national average estimates, and these estimates may vary for different geographic locations in the
country.
are estimated to be $22 billion per year in 2020
(2010$) (U.S. EPA, 2006c). Benefits for the non-road
diesel rule have been estimated at $105 billion
per year in 2030 (2010$) (U.S. EPA, 2004a). The
benefits for these and other rules with direct PM2.5
reductions are shown in Table 6-3, along with the
costs and benefit-cost ratios for each rule.1 While
the relationship between benefits and costs for PM2.5
reductions is discussed in more detail in Chapter 12,
it is useful to note that quantified benefits exceed
costs for each rule, often by a significant margin.
For the nonroad diesel rule, quantified benefits are
estimated to be over 41 times greater than costs in
2030.
As a shorthand approach for assessing potential
health benefits resulting from different mitigation
strategies when air quality modeling is unavailable,
Fann et al. (2009) developed values of monetized
health benefits per ton of emissions reduced for
SO2, NOX, and direct PM2.5 in the United States.2 For
directly emitted PM2.5 (including BC) from all sources,
these benefits (on average) range from $230,000 to
1 The benefits, costs, and associated benefit-cost ratios relate to
reductions in not only direct PM2.5 but also in other controlled co-
pollutants. EPA did not estimate the costs and benefits of controls
on direct PM2.5 or specific constituents separately.
2 The benefit-per-ton estimates found in Fann etal. (2009) reflect
a specific set of key assumptions and input data. As EPA updates
these underlying assumptions to reflect the scientific literature, the
benefit-per-ton estimates are re-estimated and are available at:
http://www.epa.gov/air/benmap/bpt.html.
$880,000 per ton of PM2.5 reduced in 2015 (2010$).3
While EPA has not separately estimated the benefits
per ton for BC reductions specifically, Table 6-4
illustrates the results for reductions in total direct
carbonaceous emissions (i.e., BC + OC) for 2015,
2020, and 2030. It is clear that controls on all sources
of direct PM2.5 can produce substantial public
health benefits in the United States; furthermore,
these benefits are 7 to 300 times greater than the
benefits-per-ton estimated for reductions of other
PM precursors such as NOX and SOX (Fann et al.,
2009), indicating that controls on direct PM2.5 may
be particularly effective for protecting public health.
The authors attribute this largely to the fact that
carbonaceous particles tend to be emitted in close
proximity to population centers.
These PM25 monetized benefit-per-ton estimates
are useful for evaluating the benefits associated
with incremental PM2.5 air quality improvements
in the United States and represent the premature
mortality and premature morbidity benefits
associated with reducing one ton of PM2 5 from a
specific source. As discussed above, these estimates
are based upon the methodology described in
3 According to Fann et al. (2009), the wide range in these benefit-
per-ton estimates reflects several key sources of heterogeneity,
including variability in source parameters which affect pollutant
dispersion and human exposure, and variability in location-specific
factors such as population density and baseline health status.
In addition, the estimates vary depending on which morbidity
and mortality effect estimate are utilized from the underlying
epidemiological references.
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Fann, et al. (2009) that used an innovative reduced-
form air quality model to estimate changes in
ambient PM2.5 concentrations resulting from a variety
of emissions control strategies applied to different
classes of emissions sources. The estimates originally
developed by Fann, et al. (2009) have been updated
to incorporate revised VSL estimates (http://www.epa.
gov/air/benmap/bpt.html). The monetized mortality
and morbidity benefits of changes in ambient PM2.5
were estimated using the Environmental Benefits
Mapping and Analysis Program (Abt Associates,
2008) and developed for specific U.S. urban areas
and the United States as a whole.4
While EPA strives to incorporate quantitative
assessments of uncertainty in the health
impacts estimates, there are aspects for which
only qualitative assessments are possible. Key
assumptions underlying the estimates are presented
in detail in the regulatory impact assessments for
each regulation. Typically, health impact assessments
include uncertainty in the concentration-response
function but are unable to include uncertainty in
emissions, simulated concentrations, and projected
population and mortality rates.
6.3.2 Global Health Benefits
Though the United States has already made great
strides toward reducing BC through its efforts to
reduce PM2.5 emissions, BC emissions remain very
high in some parts of the world due to industrial
production, residential burning of solid fuel, and
transportation (see Chapter 7). Furthermore,
unlike in the United States and Europe, where
additional controls are already planned for key
source categories such as mobile diesel engines,
emissions from many other international sources are
not yet subject to plans for control. As a result, the
largest remaining achievable increment of public
health benefits from controls on BC is international,
particularly in South and East Asia, where large
populations are exposed to high concentrations.
While a growing body of literature examines the
climate benefits of controlling BC emissions globally
(see section 6.4), only a few studies have examined
the associated health benefits. For these few studies,
the estimated public health benefits are very large,
and for many control measures the benefits greatly
exceed the costs of controls, suggesting that these
reduction measures will be advantageous for
society independent of the level of climate benefits
4 For further information about the underlying methodologies and
analytical assumptions used to develop these estimates, as well
as, relevant uncertainties involved in the estimates see Fann, et al.
(2009) and EPA regulatory impact analyses including the SO2 NAAQS
RIA (U.S. EPA, 2010h) and the Portland Cement NESHAP RIA (U.S.
EPA, 2010b) available at http://www.epa.gov/ttn/ecas/ria.htrnl.
achieved. This section (1) describes studies that have
estimated the potential health benefits that can
be achieved through mitigating BC emission and
(2) discusses approaches that have been used to
value the health benefits that could be achieved on
a global scale.
6.3.2.1 Estimating the Benefits of Global BC
Mitigation
In a study focused specifically on the health impacts
of BC reductions, Anenberg et al. (2011) estimated
that halving anthropogenic BC emissions (but not
any co-emissions) globally would avoid 157,000
(95% confidence interval, 120,000-194,000) annual
premature deaths worldwide. Over 80% of these
health benefits occurred in East Asia (China; 54%)
and South Asia (India; 31%), where large populations
are exposed to high concentrations (see blue
bars in Figure 6-1). Halving all anthropogenic BC
emissions in each major world region individually
demonstrated that the vast majority of avoided
deaths from halving BC emissions occur within the
source region, with very little impact from extra-
regional emissions. This is because BC impacts on
health are driven by surface concentrations where
humans live. BC emissions that are transported to
other regions are usually conveyed at high altitudes,
where they may have more widespread impacts on
climate, but impact human health less. Per unit of
emissions, the mortality impact of BC emissions was
estimated to be 50% larger for South Asia than for
East Asia (see red diamonds in Figure 6-1). This is
likely because emissions changes in East Asia have
smaller impacts on concentrations and because
mortality rates are higher in South Asia.
Anenberg et al. (2011) found that halving global
residential, industrial, and transportation emissions
contributed 47%, 35%, and 15% of the avoided
deaths, respectively, from halving all anthropogenic
BC emissions. Residential and industrial sector
contributions to global BC-related mortality are
each 1.2 times greater than their contributions to
global BC emissions, owing to their co-location with
dense populations, mainly in developing regions.
In contrast, the contribution of transportation
emissions to mortality is 40% lower than the
contribution of that sector to global BC emissions,
since transportation emissions are more evenly
distributed among developing and less populated
developed regions. Avoided deaths were likely
underestimated for the residential sector since
indoor exposure was excluded from the study.
Figure 6-2 shows that while the industrial and
residential sectors in East Asia have the greatest
BC emissions ("mitigation potential"), all three
sectors in South Asia have the greatest estimated
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140
SA
EU
FSU AF/ME IN
Source region
EA SE/AU
Abbreviation Region
NA
SA
EU
FSU
AF/ME
IN
EA
SE/AU
North America
South America
Europe
Former Soviet Union
Africa/Middle East
South Asia (India)
East Asia (China)
Southeast Asia/
Australia
Figure 6-1. Estimated Global Mortality Benefits of Black Carbon Reductions. Global annual avoided
premature cardiopulmonary and lung cancer deaths (thousands; blue bars) and avoided premature deaths
per Gg BC emissions reduced (red diamonds), for halving anthropogenic BC emissions in each source
region relative to the base case. (Anenberg etal., 2011)
Note:
Confidence intervals (95%) reflect uncertainty in the Concentration Response Function only.
mortality impacts per unit of emissions ("mitigation
efficiency"). Outside of South Asia and East Asia,
estimated mitigation efficiency is greatest for the
Former Soviet Union, Southeast Asia/Australia, and
Europe, while mitigation potential is likely greatest
for the residential sector in Africa/Middle East and
for the transportation sector in Europe and North
America.
It is important to note that these estimates
understate the full public health benefits that would
be achieved by reductions in global BC emissions.
Since controls to reduce BC will generally reduce
other directly emitted particles as well, halving global
BC emissions would likely result in far larger changes
in overall PM2.5 emissions. In fact, Anenberg et al.
(2011) estimated that halving global anthropogenic
OC emissions along with BC resulted in eight
times more avoided premature deaths annually
than halving BC alone. Nevertheless, this study
demonstrates that BC mitigation efforts are likely
to be more effective at reducing mortality in some
regions than others, largely driven by population
exposure. Although the coarse grid resolution
(~170 km on a side) used by Anenberg et al. (2011)
was unable to capture fine-scale spatial gradients
in population and concentration, emissions from
different sectors result in different exposure patterns.
Therefore, the health response to controlling
emissions from different regions and from different
source sectors is likely to vary. Finer scale models
can be used to investigate how different mitigation
strategies impact health within individual world
regions.
While Anenberg et al. (2011) examined broad
percentage decreases in BC emissions from
individual source regions and sectors, actual
mitigation measures will affect the full mixture of
emissions from individual sources. Currently, the
most comprehensive assessment of more realistic
emissions control measures is the Integrated
Assessment of Black Carbon and Tropospheric Ozone
sponsored by the United Nations Environment
Programme (UNEP) and the World Meteorological
Organization (WMO). Using an integrated modeling
approach addressing a range of co-emitted
pollutants, the UNEP/WMO Integrated Assessment
identified a small number of emissions reduction
measures that would achieve major benefits for
near-term climate change. This suite of measures
included both BC reduction measures and methane
reduction measures. For BC, the assessment
modeled the impact of both "technical measures,"
such as improving coke ovens and brick kilns and
increasing use of diesel particulate filters, and
"non-technical measures," such as eliminating
high-emitting vehicles, banning open burning
of agricultural waste, and eliminating biomass
cookstoves in developing countries. Specifically, the
Assessment evaluated the health benefits of the
following BC measures:
• Use of diesel particle filters as part of a Euro VI
package for on-road and off-road diesel vehicles
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1000
& 100
o 10
A
•
1 Uncertainty in
mortality estimate
^
• A
• +
*
•
Uncert
BCem
D Residential
A Industrial
A°
O
aintyin
ssions
^ Transportation
Receptor region;
• NA AF/ME
• SA • IN
EU • EA
• FSU • SE/AU
D global
10
100
1000
10000
BC emissions (Gg)
Figure 6-2. Annual Avoided Premature Cardiopulmonary and Lung Cancer
Deaths Per Unit BC Emissions Reduced ("mitigation efficiency") versus Total
BC Emissions (Gg; "mitigation potential") for Particular Source Sectors within
Each Region. (Anenberg et al., 2011)
Notes:
1. Avoided deaths are estimated in the three simulations where global emissions in each
sector are halved, and shown for each receptor region; these deaths are compared with
emissions from each region, assuming that deaths from inter-regional transport are
negligible.
2. Uncertainty in the mortality estimates is calculated from the uncertainty in the CRF only
(-22% and 56% from mean for cardiopulmonary and lung cancer mortality).
3. Uncertainty in BC emissions is assumed to be a factor of 2 from the central estimate
(Bond et al., 2004; 2007).
4. Since these uncertainties are factor differences from the central estimate, they are
identical for each data point.
• Introduction of clean-burning cook stoves for
cooking and heating in developing countries
• Replacement of traditional brick kilns with vertical
shaft kilns and Hoffman kilns
• Elimination of high-emitting vehicles in on-road
and off-road transport (excluding shipping)
• Ban of open field burning of agricultural waste
• Substitution of clean-burning cook stoves using
modern fuels for traditional biomass cook stoves
in developing countries
Together, these BC measures were estimated to
reduce global anthropogenic BC emissions by
75%, along with substantial
reductions in co-emitted OC,
NOX, and CO.
The UNEP/WMO Assessment
estimated that fully
implementing these measures
by 2030 would avoid 0.6-
4.4 million PM2.5 related
premature deaths and 0.04-
0.52 million ozone-related
premature deaths annually
around the world, based on
2030 population projections
(Shindelletal., 2012; UNEP
and WMO, 2011b). Consistent
with the results of Anenberg
eta I. (2011), over 80% of
the health benefits occur in
Asia. Figure 6-3 shows that
implementing the BC and
methane measures would
reverse the trend of increasing
air pollution-related deaths
in Africa and South, West,
and Central Asia (although
methane and BC measures
are shown together here, BC
measures contribute -98%
of the total health benefits).
Figure 6-3 also shows the
additional benefits achievable
in areas already making
progress. The study also found
that the substantial health
benefits of the joint air quality/
climate mitigation measures
examined occur regardless of
whether measures to reduce
long-lived GHG have been
implemented. A follow-on
study to the UNEP/WMO Assessment found that in
Africa, Asia, and Latin America and the Caribbean,
improved biomass cookstoves would generate
the greatest health benefit of all the measures
examined, with substantial additional benefits
from mitigation measures for the transportation
sector (UNEP, 2011). In Europe and North America,
switching to pellet stoves from current domestic
wood-burning technologies was estimated to
deliver the greatest health benefit. The study also
found that banning the burning of agricultural
crop residues would produce a small benefit in all
regions.
One additional study examined the potential
health benefits of global reductions in vehicle
emissions specifically, accounting for the full
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1.5
0.5
0
-0.5
-1
-1.5
-2
-2.5
-3
Africa
Northeast
Asia,
Southeast
Asia and
Pacific
Latin North
America America
and and
Caribbean Europe
Reference
Reference +
CH4+BC
measures
South,
West and
Central Asia
Figure 6-3. Comparison
of Premature Mortality
by Region (millions
of premature deaths
annually), showing
the change in 2030
in comparison with
2005 for the reference
scenario emission
trends and the reference
plus methane and BC
measures. The lines on
each bar show the range
of estimates. Over 98%
of the benefits were
attributed to the BC
measures. (Source: UNEP
and WMO, 2011 b)
mixture of emissions affected by control measures
(Shindell et al., 2011). Emissions from on-road motor
vehicles, including cars, trucks, and motorcycles,
are growing rapidly in many countries, due to rising
personal vehicle ownership and usage. Shindell et
al. (2011) assessed the potential health benefits of
imposing existing gasoline and diesel European
vehicle emissions standards in developing regions.
Specifically, the study examined the adoption of Euro
6 vehicle standards in China and India, Euro 4 vehicle
and Euro 3 motorcycle standards in Africa and the
Middle East, Euro 6 vehicle and Euro 3 motorcycle
standards in Brazil, and Euro 3 motorcycle
standards in the rest of Latin America, based on
the authors'judgment of local financial, technical,
and institutional capacity. The European standards
are more stringent than the standards currently
planned in these regions. Imposing these standards
was estimated to avoid 120,000-280,000 premature
PM2.5 and ozone-related deaths worldwide, based
on 2030 population, largely resulting from local
emissions controls. Small benefits were also seen in
areas with no additional local emission controls due
to long-range transport of ozone and PM2.5 in the
atmosphere. The study concluded that tighter vehicle
emissions standards are likely to lead to significant
health benefits, in addition to climate benefits in
most cases.
In addition to the BC and sector-specific studies
described above, there is a small but emerging
body of literature assessing the global health
benefits of PM2.5 emissions reductions. The results
from additional global and international studies are
summarized in Appendix 3. Many of these studies
estimate the avoided premature deaths associated
with reductions in BC and other PM2.5constituents,
while other studies attempt to compare the costs
and benefits of potential mitigation strategies. These
studies indicate that a large number of premature
deaths can be avoided annually by undertaking
strategies to reduce BC emissions (Shindell et al.,
2011; 2012; Anenberg et al., 2011; Wilkinson et al.,
2009; Saikawa et al., 2009; Jacobson, 2010). The
studies that include a benefit-cost comparison show
that estimated human health benefits significantly
exceed the estimated costs for certain BC mitigation
strategies (Smith and Haigler, 2008; Kandlikar et
al., 2009; Baron et al., 2009). Thus, BC reductions
appear advantageous to society independent of the
level of climate benefits achieved. This is particularly
true of sources associated with high human health
exposures, such as cookstoves (which are often used
indoors in confined spaces) and vehicles located in
densely populated areas.
Valuing the global health benefits of emissions
controls is difficult due to limited data on willingness
to pay for reducing mortality risks. In general,
studies valuing reduced mortality around the world
have used two methodologies. In the first, a uniform
VSL, generally from U.S. or European studies, is
applied to avoided premature deaths in all countries,
regardless of income or other economic disparities.
While ethically appealing, willingness-to-pay is a
function of income and thus would reasonably be
expected to vary around the world. The second
approach is to adjust VSLs from the developed world
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by measures of income or other economic welfare
in other countries, for example income elasticity.
While this method is often criticized because it
seems to imply that lives in different countries are
valued at different levels, it accounts for differences
in income levels and is therefore often the preferred
approach among economists. This approach still has
limitations, including the inability to account for
differences in social values and culture.
Acknowledging the advantages and disadvantages
of these valuation approaches, the UNEP/WMO
Assessment and Shindell et al. (2011) reported the
monetized health benefits of mitigation measures
using both methods (UNEP and WMO, 2011a). The
UNEP Assessment estimated the monetary value
of the ozone and PM2.5-related premature deaths
avoided as a result of the methane and BC measures
was $1.7-10.9 trillion in 2010 $US. Consistent with the
mortality results, the vast majority of the monetized
health benefits resulted from the BC measures.
Shindell et al. (2011) estimated that the global
health benefits of imposing tighter vehicle emissions
standards in the developing world are valued at
$0.7-2.6 trillion in 2010 $US. Since the majority of
the health benefits resulting from these emissions
control measures occur in developing countries,
the income-adjusted VSL approach leads to lower
valuation estimates compared with the uniform
VSL approach.5 Shindell et al. (2012) conclude that
since about half of the benefits of all BC mitigation
measures are attributable to improved efficiencies
for implementing improved brick kilns and cleaner
burning stoves, which lead to net cost savings,
and another 25% to regulatory measures on high-
emitting vehicles and banning agricultural waste
burning, which require primarily political rather
than economic investment, the majority of the BC
measures could be implemented with substantially
greater benefits than costs.
6.3.3 Conclusions Regarding Potential
Health Benefits
All control measures that reduce PM2.5 pollution
are virtually certain to achieve health benefits.
Programs aimed at reducing PM2.5 in the United
States, such as rules targeting light and heavy duty
vehicles, diesel emissions, and marine vessels, as
well as industrial stationary sources, have greatly
reduced PM2.5 concentrations (including BC) and
PM25-related mortality. These programs have very
favorable benefit-cost ratios, particularly for the
mobile source sector. While progress has been
5 The UNEP/WMO Assessment (2011) did not evaluate the full
costs of implementing the modeled measures, but did report cost
information where available for key demonstration projects in
different countries and sectors.
made, the PM2.5 health burden in the United States
remains significant. EPA has introduced a number
of programs for both mobile and stationary sources
that are estimated to have a substantial impact
on air quality and, as a result, PM25-related health
impacts, over the next several decades. However,
additional controls for transportation and stationary
sources, as well as for residential wood burning,
can further reduce the remaining BC emissions
(Chapters 8-11).
The largest opportunities for achieving the health
benefits of BC mitigation measures are in lesser
developed countries due to high emissions located
in densely populated areas, particularly in South and
East Asia. Although the body of literature is limited,
available studies demonstrate that mitigating BC
emissions would have substantial benefits for
global public health, potentially avoiding millions of
premature deaths each year valued in the trillions
of $US. Although valuing health benefits around
the world is complicated by data limitations, several
studies undertaking such analyses have found that
the mortality benefits alone are quite substantial
and may alone justify mitigation efforts. Reducing
BC emissions from transportation and residential
sources, in addition to some BC-rich industrial
sources, would likely achieve the greatest combined
health and climate benefits. More information on
the benefits and costs of individual measures in each
country is needed to support policy decisions made
at the national level.
6.4 Climate Benefits of Reducing
Black Carbon Emissions
A number of recent studies and assessments have
pointed to the possibility that reducing BC could
provide climate benefits within the next several
decades. Some of these studies have focused
exclusively on BC, without adequately treating co-
emitted pollutants, and/or have estimated direct
forcing effects only, without accounting for the
potential off-setting cloud interaction effects. As
the treatment of BC's atmospheric chemistry and
co-pollutants in climate models has advanced,
however, studies have begun to focus on certain
key sectors and regions as potentially fruitful
mitigation options for climate. The recent UNEP/
WMO (2011a) assessment, for example, indicates
that a small group of carefully targeted BC measures
could help improve chances of keeping the Earth's
temperature increase to less than 2°C relative to
pre-industrial levels (see section 6.4.1). However, the
climate benefits of reducing BC emissions are less
well understood and less certain than the public
health benefits. Because BC concentrations and
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their climate impacts vary spatially and temporally
(as discussed in Chapter 2), the location and timing
of emissions reductions is critically important for
estimating climate benefits of mitigation. In addition,
emissions control measures for BC also reduce co-
emitted pollutants that lead to cooling (e.g., SO2,
OC). Because many of these co-emitted pollutants
lead to climate cooling, the climate benefits of BC
emissions control measures may be offset. Therefore,
the full mixture of emissions must be considered
in estimating the climate benefits of potential
mitigation measures. Since these factors are complex
and often not well understood, quantitative analysis
of the climate benefits of BC mitigation strategies is
difficult and the number of related studies is limited.
Of the studies currently available, some have focused
on the physical climate benefits of BC mitigation—
estimating changes in temperature, ice melt, or
radiative forcing. A number of these studies explicitly
compare the climate benefits of BC reductions to
the climate benefits of reductions in other GHGs.
These studies often use metrics such as GWP or
GTP (introduced in Chapter 2) as the basis for
comparing alternative climate mitigation strategies.
Other studies have extended the analysis of climate
benefits by attempting to place an economic value
on avoided impacts. In a few cases the economic
benefits of particular BC mitigation strategies
were compared to those of alternative strategies
targeting either BC or long-lived GHGs. The next
several sections describe (1) the potential physical
benefits that can be achieved through BC mitigation,
(2) how those benefits compare to benefits that
could be achieved through CO2 mitigation, and
(3) the potential value of the climate benefits of BC
mitigation.
6.4.1 Studies Estimating Physical Climate
Benefits
As discussed in Chapter 2, the nature and
distribution of BC and its mechanisms of action mean
it can have important direct and indirect effects on
climate that differ from those of GHGs. Unlike with
GHGs, these effects are not limited to those derived
from radiative forcing on a global scale. Rather,
the effects associated with BC include alteration of
cloud properties, which affects cloud reflectivity,
precipitation, and surface dimming. In addition,
deposited BC can result in disproportionate warming
in areas covered by snow and ice, which is greatest
near source regions (e.g. the Himalayas) but still
significant in the Arctic.
Most studies on the climate benefits of BC mitigation
have focused on estimating the impacts of broad-
scale global or regional emissions reductions.
Several investigators have used global climate
models to examine reductions of BC, OC and
in some cases associated GHGs from the fossil,
biofuel, and biomass sector sources. Most of these
have focused on the effect of global reductions
on radiative forcing or temperature. As discussed
below, the results generally suggest that the
largest climate benefits are likely to accrue from
strategies that reduce emissions from BC-rich
sectors such as mobile diesel engines and other
fossil fuel combustion sources, as opposed to
sectors where the quantity of BC compared to co-
emitted pollutants is smaller (e.g., biomass burning).
Below, we discuss (1) the global climate benefits
of BC emissions reductions, (2) benefits of BC
reductions specific to ice-covered regions, and (3)
key uncertainties in these estimates that are due to
insufficient scientific understanding of how aerosols
affect climate.
6.4.1.1 Global Climate Benefits of BC Reductions
As discussed previously, some co-emitted pollutants
lead to cooling that can counteract the warming
by BC. Changes to the entire emissions mixture
must therefore be considered when estimating
the climate impacts of BC mitigation measures.
Several studies have examined the climate impacts
of eliminating all emissions from individual sources
and found the largest and most consistent benefits
in terms of negative forcing (cooling) result from
reductions in emissions from fossil fuel sources. For
example, Jacobson (2010) found that eliminating
all fossil fuel soot reduced surface air temperature
by 0.3-0.5 K (13-16% of total net global warming).
Another study that evaluated multiple models found
that global reductions in open biomass burning
(where the emissions mixture typically includes
a higher concentration of cooling compounds)
produced small but positive climate benefits (Kopp
and Mauzerall, 2010).
Several studies have examined the climate benefits
of emissions reductions from individual emissions
source sectors, again finding that sectors with higher
BC ratios generally have larger positive forcing.
These studies are generally consistent in finding that
transportation and household biofuel combustion
contribute more than any other sector to positive
forcing. For example, Unger et al. (2010) examined
the warming impacts of each major sector's
emissions, taking into account the full mixture of
aerosols and gases (short-lived and long-lived) from
each sector. On-road motor vehicles and household
biofuels, major sources of global anthropogenic
BC emissions, were found to contribute more than
any other sector to globally averaged near-term
warming (by 2020) (Figure 6-4). Koch et al. (2007a)
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Chapter 6
(3) On-road(199)
Household biofuel (132)
Animal Husbandry (98)
Household fossil fuel (84)
Waste/landfill (84)
Power (79)
Agriculture (29)
Off-road land (20)
Aviation (-6)
Agr. maste burning (-14)
Shipping (-43)
Industry (-1 58)
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•
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1
• Ozone
Sulfate
• Nitrate
• Black carbon
• Organic carbon
AIE
• Methane
Nitrous Oxide
• Carbon Dioxide
-400 -200 0 200
Radiative forcing (mWrrr2)
( D I
On-road(417)
Household fossil fuel (254)
Household biofuel (159)
Animal Husbandry (131)
Agriculture (98)
Waste/landfill (88)
Off-road land (39)
Aviation (27)
Agr. waste burning (-14)
Shipping (-22)
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i
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i
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Sulfate
Nitrate
Black carbon
Organic carbon
AIE
Methane
Nitrous oxide
Carbon dioxide
-600 -400 -200 0 200 400 600
Radiative forcing (inWnr2)
Figure 6-4. Global Radiative Forcing Due to Perpetual
Constant Year 2000 Emissions, Grouped by Sector, in
2020 (Top) and 2100 (Bottom) showing contribution from
each species. The sum is shown on the title of each bar,
with a positive radiative forcing indicating that removal of
this emissions source will result in cooling. From Linger et
al. (2010) (also shown in Chapter 2).
examined the direct radiative forcing of aerosols
only, finding that the sectors that are responsible
for the largest BC radiative forcing are residential
(0.09 W m 2) and transport (0.06 W rrr2), but that
co-emitted scattering (i.e., cooling) components
reduce these impacts to 0.04 W m 2 and 0.03 W
m"2, respectively. Shindell et al. (2008a) found that
across-the-board emission reductions in household
fuel burning in Asia and in transportation in North
America are likely to offer the greatest potential for
near-term climate benefits. Although these studies
were limited to direct radiative effects, Bauer et al.
(2010) also found that reducing diesel emissions
would reduce positive forcing (i.e., warming) even
when accounting for cloud changes, which can
result in cooling. In another study examining the
effects of all aerosols, Bauer and Menon (2012)
reached similar conclusions. This study focused on
regional differences in the impact of emissions from
different source categories, and concluded that
the largest opportunities to reduce positive forcing
due to all aerosols included transportation in all
regions, agricultural burning in Europe and Asia, and
residential cooking and heating ("domestic sector")
in Asia.
This body of literature suggests that both
transportation and residential BC sources may
be attractive targets for BC mitigation measures.
Although several of these studies found that
reductions in industrial and power generation
emissions may accelerate near-term warming
(Unger et al., 2009; Shindell et al., 2008a), this broad
categorization includes a variety of different types
of sources, some of which are major emitters of SO2,
a precursor of sulfate (a "cooling" aerosol). Within
the industrial sector, however, are some sources
(e.g., brick kilns, coke ovens) that are major emitters
of BC in the developing world. Controlling emissions
from these specific BC-rich sources will likely also
lead to climate benefits, as discussed below. As
with any strategy development, determining the
specific optimal measures to implement depends on
a number of factors in addition to the climate and
public health benefits. These factors are discussed in
more detail in Chapter 7.
Several recent studies have examined how specific
emission control measures are expected to reduce
emissions. To date, the most comprehensive
assessment of this type is the UNEP/WMO
Assessment described in Section 6.3.2 (UNEP and
WMO, 2011a; Shindell et al., 2012). This study
found that implementing an illustrative set of BC
and methane (CH4) emission control measures
together would reduce future global warming by
0.5°C (0.2°C - 0.7°C), with about half the reduction
specifically from the BC measures. Implementing
the BC and CH4emission control measures by 2030
was estimated to halve the expected increase in
temperatures for 2050 compared with the reference
scenario (based on current policies and energy and
fuel projections), as shown in Figure 6-5. This study
used a range of values from the literature reflecting
the indirect and direct radiative forcing effects of BC
and OC to provide a range of expected outcomes
that account for uncertainty. A follow-on study to
the UNEP/WMO Assessment found that the measure
likely to produce the greatest near-term global
climate benefit is switching from traditional biomass
cookstoves to cleaner burning stoves, followed by
reducing emissions from the transportation sector
(UNEP, 2011).
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Because transportation emissions are expected to
contribute the most to BC-related warming in the
future (Koch et al., 2007b), Shindell et al. (2011)
examined the climate benefits of imposing tighter
vehicle emission standards in China, India, Africa,
the Middle East, and Latin America. Relative to no
additional controls, imposing these standards led
to significant reduction in warming in the Northern
Hemisphere extra-tropical region (reduction of
0.22°C, with a potential range from 0.04°C to
0.38°C) and in the Arctic (reduction of 0.28°C, with
a potential range from 0.02°C to 0.47°C) over the
next 50 years, although the total reduction in global
warming after 2040 overall was small. Controlling
emissions from heavy-duty diesel trucks in India
and Brazil was found to have the greatest climate
benefits, followed by controls on medium-duty
diesel vehicles in India and light-duty petrol vehicles
in North Africa and the Middle East. Controlling
emissions from light-duty gasoline vehicles
everywhere and from motorcycles and medium-
duty trucks in some regions also provides climate
benefits that are more limited. These BC reductions
were also associated with precipitation changes,
but such results are highly uncertain and warrant
further study. Despite large uncertainties, this study
demonstrates the substantial climate benefits of
controlling emissions from the motor vehicle sector
around the world.
Few studies have examined the climate benefits
of specific particle control programs on smaller,
more localized scales. A recent study of particular
relevance examined the results from California's
laws to reduce particle pollution, in particular those
regulating diesel emissions. The study found that
these rules reduced atmospheric concentrations of
BC with a measurable impact on regional radiative
forcing. Modeled results indicate that the decrease
in BC emissions in California has led to a cooling
cr>
CO
2
cu
2
O)
Q.
£
Reference
CO, measures
CH, + BC measures
BC measures
1900
1950
2000
2050
Figure 6-5. Observed Deviation of Temperature to 2009 and Projections under Various Scenarios.
Immediate implementation of the identified BCandCH4 measures, together with measures to reduce CO2
emissions, would greatly improve the chances of keeping the Earth's temperature increase to less than
2°C relative to pre-industrial levels. The bulk of the benefits of CH4 and BC measures are realized by 2040.
(UNEPandWMO,2011a)
Note: Actual mean temperature observations through 2009, and projected under various scenarios thereafter, are shown
relative to the 1890-1910 mean temperature. Estimated ranges for 2070 are shown in the bars on the right. A portion of
the uncertainty is common to all scenarios, so that overlapping ranges do not mean there is no difference. For example, if
climate sensitivity is large, it is large regardless of the scenario, so temperatures in all scenarios would be towards the high
end of their ranges.
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Chapter 6
of 1.4 W rrr2 (±60%) (Bahadur et al., 2011). So,
while uncertainties remain, as outlined in previous
chapters, emerging research suggests that targeting
emissions reductions from key sectors can have
measurable benefits for climate.
6.4.1.2 Benefits of BC Reductions on Snow- and Ice-
Covered Regions
As noted above, BCs effects are highly regionalized
and include impacts on precipitation, atmospheric
stability, and snow/ice melting that differ in
important ways from those driven by GHGs. Global
modeling provides useful insights into potential
responses in regions identified in Chapter 2 as being
particularly affected by BC emissions. For example,
Jacobson (2010) found the extreme strategy of
eliminating all anthropogenic emissions from sources
of fossil fuel and biofuel BC would reduce global
temperatures by 0.4 to 0.7° C; with a reduction of
about 1.7° C in the Arctic. This is consistent with
other modeling and analysis discussed in Chapter 2
that suggest a larger impact of BC and other
pollutants on the Arctic, and thus greater potential
benefits from emissions control measures.
Recent findings of the Arctic Council Task Force on
Short-Lived Climate Forcers suggest that mitigating
sources of BC emissions in or near the Arctic will
have greater climate benefits in that region, with
important seasonal and spatial variations. Impacts
in the Arctic are greatest during the spring and
summer months when the solar radiation is the
strongest. Specifically, in its 2011 Progress Report and
Recommendations for Ministers (Arctic Council, 2011),
the Task Force noted that:
[l]n the Arctic, the potential for... offsetting
effects from non-black carbon aerosols is weaker.
Over highly reflective surfaces such as ice and
snow in the Arctic, the same substances that
might cool the climate in other regions may cause
warming since they are still darker than ice and
snow. This warming impact is magnified when
black carbon physically deposits on snow or ice.
Emissions closer to the Arctic have a greater
chance of depositing, and thus appear to have
greater impact per unit of emission.
The Task Force highlighted the importance of
reducing emissions from in-Arctic sectors such as
land-based transportation, open biomass burning,
residential heating, and marine shipping in the
Arctic, but also noted that emissions from outside
the Arctic, especially those in close proximity to
the Arctic, are important for Arctic climate change,
partly because of the volume of these emissions
and (as noted above) because of the relatively small
cooling effect of co-emitted pollutants in the region.
The 2011 Arctic Monitoring and Assessment
Programme (AMAP) report, "The Impact of Black
Carbon on the Arctic" examined the influence of BC
emitted from different sources and world regions
on radiative forcing and temperature in the Arctic
(Quinn et al., 2011). This study found that in general,
BC deposition on snow and ice in the Arctic (which
contributes to the snow/ice albedo effect described
in section 2.6) exerts a greater warming effect than
within-Arctic direct radiative forcing from BC in the
atmosphere. Furthermore, both direct forcing and
snow/ice albedo forcing in the Arctic per unit of
emissions were shown to be highest for emissions
originating near to or within the Arctic region, with
RF per unit BC emission, mW/m2 per Gg/y
2.0 r
1.5
Normalized net
forcing for Arctic Council
Nation emissions
Normalized net
forcing for emissions
by latitude band
Normalized net
forcing for projected
shipping emissions
Within
Arctic
Figure 6-6. Summary of Normalized Net Forcing per Unit of Emissions (includes atmospheric direct
forcing by BC and BC-snow/ice forcing) due to emissions from Arctic Council nations and the rest of the
world, the indicated latitude bands, and global and within-Arctic shipping. (Source: Quinn et al., 2011)
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the Nordic countries (i.e., highest latitudes) having
the highest forcings per unit of emissions, followed
by Russia, Canada, and the United States (Quinn
et al., 2011, p. 95). These impacts are illustrated in
Figure 6-6. this is because within-Arctic BC sources
are more likely to cause warming near the surface
and to lead to BC deposition on snow and ice
surfaces. Furthermore, warming at high altitudes
reduces atmospheric energy transport into the
Arctic; as a result, AMAP (Quinn et al., 2011) found
that direct atmospheric forcing by BC that has been
transported into the Arctic at high altitudes may
have a relatively small impact on Arctic surface
temperatures.
However, the Arctic climate is coupled with that
of the Northern Hemisphere and is thus sensitive
to changes in radiative forcings in other nearby
regions. Since the bulk of global BC emissions occur
outside the Arctic, and since global BC forcing
results in poleward transfer of heat energy, Arctic
temperatures are significantly affected by BC
direct forcing occurring outside the Arctic region.
Overall, when total global emissions are considered,
BC emissions in the rest of the world are the
dominant influence on radiative forcing in the Arctic
(Figure 6-7). The AMAP results confirm both that
latitude and the total magnitude of emissions matter:
It has been suggested that emissions north of40°N
have a large impact on the Arctic particularly
in the winter and spring when the polar dome
extends to the mid-latitudes over Europe and Asia
.... To test this assumption and to compare the
potential impact of sources on Arctic climate as
a function of latitude between 40°N and 90°N, a
set of experiments was performed with emissions
gridded by latitude band. The latitude bands
included in the analysis were 90°S to 40°N, 40°N
to 50°N, 50°N to 60°N, and 60°N to 90°N....
[Ejmissions in the most southerly latitude band
(90°S to 40°N) result in the largest direct RF due
to the magnitude of the emissions of BC in the
northern hemisphere tropics and mid-latitudes.
Atmospheric direct RF also is relatively large
for the 40°N to 60°N latitude band because of
the magnitude of emissions and likelihood of
transport to the Arctic. Emissions within the Arctic
(60°N to 90°N) result in a smaller absolute direct
RF because of their lower magnitude compared
to emissions in more southerly latitude bands.
Absolute BC-snow/ice RF increases with latitude
between 40°N and 60°N. This result confirms that
emissions from lower latitudes are less effectively
deposited in the Arctic since they reach the Arctic
at higher altitudes .... Normalized BC-snow/ice
RF increases dramatically with increasing latitude
band... confirming the efficiency with which
Atmospheric direct (BC+OC)
RF, mW/m2
(a) 70
60
50
40
30
20
10
0
RF per unit BC emission, mW/m2 per Gg/y
(b) 0.25 r
0.20
0.15
0.10
0.05
60° N to 90° N
60° N to 90° N
90°S-40°N 40°-50°N 50°-60°N
Source region
60°-90°N
Figure 6-7. Contribution to Radiative Forcing of
Carbonaceous Aerosol Emissions within Different
Latitude Bands, (a) Absolute and (b) Normalized per
unit emission atmospheric direct radiative forcing due to
BC+OC and BC-snow/ice radiative forcing as a function of
latitude band. (Quinn et al., 2011, Figure 8-9)
sources close to the Arctic are transported to
and deposited within the Arctic. This result also
indicates that per unit emission, sources within
the Arctic yield the largest RF. (Quinn et al., 2011,
p. 57)
Importantly, AMAP (Quinn et al., 2011) indicates that
reductions in OC emissions will provide benefits
in the Arctic too, since OC emissions appear to
exert positive radiative forcing over snow- and ice-
covered surfaces. Note that the radiative forcings
reported in Figure 6-7 are for the combination of
OC+BC emissions. Similarly, sulfate aerosols, which
normally exert a cooling influence, appear to have
a much weaker cooling effect over snow and ice
(Quinn et al., 2011). While aerosol indirect forcing
effects in the Arctic are highly uncertain, current
evidence suggests that indirect and semi-direct
effects are less negative in the Arctic than for the
global average. This all means that a wide array of
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Chapter 6
measures aimed at reducing BC and OC in or near
the Arctic will provide benefits to the Arctic region,
as will BC-focused strategies elsewhere. AMAP
(Quinn et al., 2011) concluded that global strategies
to manage BC emissions are necessary to mitigate
Arctic climate change.
There is relatively less information regarding
specific benefits of BC reductions for other snow-
and ice-covered regions, including the Himalayas.
As discussed in section 2.6.5, Menon et al. (2010)
modeled the impacts of estimated increases in
BC between 1990 and 2000 and found that in
particular BC from fossil fuel/biofuel use in India
may be responsible for some of the observed
patterns and trends in snow and ice melting and
precipitation in the Himalayan region. Such changes
may have significant implications for water supply
in the region. While a number of studies have
suggested BC and associated emissions may play
a role in reduced monsoon rains, current modeling
capabilities do not provide a basis for reliable
quantitative assessments of the extent to which
emissions reductions might reverse observed
changes in precipitation.
6.4.1.3 Key Uncertainties in Estimating Climate
Benefits of BC Reductions
As discussed in detail in Chapter 2, the results from
the studies described in the previous subsections
have some level of uncertainty. The primary sources
of this uncertainty include lack of understanding
about:
• The climate effects of reductions in co-emitted
pollutants, especially brown carbon (BrC)
emissions and the extent to which they contribute
to radiation absorption (Magi, 2009).
• The effect of model representation of aerosol
mixing state and aerosol-cloud interactions,
including radiative effects and precipitation
effects, on estimated climate impacts. These
processes can have a major influence on the
overall warming or cooling effect of emissions
changes.
• Effects of non-BC aerosols in the Arctic.
• Effects of other atmospheric processes (such as
atmospheric transport and deposition) on climate
outcomes.
In addition, errors in the emissions inventories of BC
and OC and other reflecting agents from each sector,
particularly for residential solid fuel combustion, may
lead to over or under estimation of the magnitude
of the climate impact.
6.4.2 Comparing Climate Benefits of
Reductions in BC vs. CO2
While studies performed to date do not include the
full set of aerosol interaction effects, co-emissions,
or other uncertainties, taken as a whole they
suggest that reductions of BC may be among the
most effective strategies for reducing near-term
warming, and can complement GHG reductions
as part of an overall climate strategy (Grieshop et
al., 2009; Kopp and Mauzerall, 2010). As described
in Chapter 2, BC reductions can reduce the rate
of climate change and provide climate benefits in
the near term. However, BC reductions today have
much smaller effects on temperatures in 100 years.
Therefore, BC emissions reductions cannot substitute
for CO2 reductions for purposes of alleviating long-
term warming. Studies indicate that BC emissions
reductions that come at the expense of reductions
in CO2 emissions would result in short-term cooling
but add an additional commitment to long-term
radiative forcing due to the life time of CO2 in the
atmosphere (Lack et al., 2009).
The UNEP/WMO BC and Tropospheric Ozone
Assessment (UNEP and WMO, 2011a) described
in sections 6.3.2 and 6.4.1 compared the climate
benefits of groups of BC and CH4 mitigation
measures to a scenario developed by the
International Energy Agency in which long-lived
GHG concentrations were reduced to a level of
450 ppm CO2eq (International Energy Agency, 2009).
As illustrated by Figure 6-5, the reductions in CH4
and BC combine to produce a noticeable impact on
near-term warming as compared to the reference
case or CO2 measures by themselves. The analysis
showed that even aggressive CO2 reductions may
not keep climate change from approaching 2°C by
mid-century.6 At the same time, it is important to
note that the benefits of reducing BC and CH4 are
insufficient to avert warming over the long term.
Reducing short-lived climate forcers now, while
neglecting to achieve aggressive CO2 reductions,
may not keep temperatures from reaching the
2°C mark in 2070 and beyond. These results, and
those from other studies on the temporal aspects
of reducing BC and other short-lived forcers,
underscore the scientific rationale for reducing long-
lived GHGs and BC simultaneously as two distinct,
complementary strategies that act on different time
scales to address global warming and other effects
of climate change.
6 An increase in global mean temperatures of 2°C since
preindustrial times was adopted as an international target under
the UN's Copenhagen Accord in December 2009.
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Diesel, modern
Benefits of Reducing Black Carbon Emissions
Diesel, superemitter Biofuel, traditional cooking stove
0.5 1 1.5 2 2.5
Total energy added during lifetime (GJ)
10 20 30 40
Total energy added during lifetime (GJ)
012345
Total energy added during lifetime (GJ)
Figure 6-8. Integrated Forcing by Aerosols Emitted from Burning 1 Kg of Fuel from Different Sources,
based on results of 250 Monte Carlo simulations. (Note scale differences.) (Bond, 2007, Figure 6)
Among the other studies comparing BC and CO2
strategies, Grieshop et al. (2009) used a valuation
that "one ton of black carbon causes about 600
times the warming of one ton of carbon dioxide
over a period of 100 years" to state that eliminating
present-day emissions of BC over the next 50 years
would have "an approximately equivalent climate
mitigation effect to removing 25 Gt C from the
atmosphere over the same period."7
Some studies also account for changes in co-
emissions from BC and GHGs that may affect
climate outcomes. For BC, the co-emissions (e.g.
SO2, OC) are often reduced by the same control
measures that affect BC and the climate impacts
are generally realized on the same timescale as (i.e.,
short atmospheric lifetimes lead to near-term climate
effects). For CO2, there is a gap between climate
effects due to CO2 and effects due to co-emitted
species, since the latter generally have a shorter
lifetime. Furthermore, with CO2, it is often possible
to reduce co-emissions separately with end-of-
pipe technologies, which makes it possible to make
independent decisions about how much to reduce
CO2 vs. co-emissions from a given source. All of this
illustrates the importance of accounting for co-
emissions in climate models, but also the complexity
of assessing the impact of specific control measures.
Other studies have indicated that implementing
aerosol mitigation measures for BC-rich sources
can yield more cooling over the short term (10-20
years) than eliminating CO2 emissions from those
sources (Bond, 2007; Jacobson, 2005; Sarofim,
7 In this study, 25 Gt was chosen because it equals one "wedge"
from the Pacala and Socolow (2004) study that identified large-scale
mitigation options over the next 50 years. However, Grieshop et al.
(2009) did not involve any calculations to compare the short-term
and long-term effects of implementing a BC wedge rather than an
additional greenhouse gas wedge, did not examine co-emissions,
and did not take into account cloud interaction effects.
2010). Sarofim (2010), for example, addressed one
specific mitigation option (retrofits of some U.S.
diesel vehicles) and showed that the CO2 equivalent
reductions calculated by using a GWP would lead
to radiative forcing reduction from black carbon
mitigation peaking in the year that the vehicles are
retrofitted and dropping to almost zero change
in global radiative forcing after 20 years as the
retrofitted vehicles are retired. In contrast, the
radiative forcing reduction from the CO2 equivalent
mitigation calculated using GWPs peaks about a
decade after the start of the mitigation period at
only a tenth of the BC peak, but at the end of the
century the radiative forcing reduction is still more
than half of what it was at that peak.
Bond (2007) examined emissions from multiple
source types and compared the total radiative
(integrated) forcing from those sources over 20
years for carbonaceous aerosols (both OC and BC)
to the integrated forcing from CO2 (an approach
similar conceptually to using GWP weightings). The
study showed that the aerosol emissions resulting
from burning 1 kg of fuel in a super-emitting diesel
vehicle has more than a 90% chance of contributing
more total forcing than CO2 from that source over a
20 year timeframe, and even for a normal (pre-2007)
diesel, the aerosol emissions resulting from burning
1 kg of fuel are likely to contribute more than half as
much warming as the CO2 emissions over 20 years
(see Figure 6-8). This study did not account for the
indirect effects of aerosols or snow albedo effects.
Jacobson (2005) did include co-emissions and more
cloud interactions, and still found that diesel vehicles
warmed climate more than gasoline vehicles for
13-54 years, because the higher BC emissions from
diesel vehicles outweighed the lower CO2 emissions
over that timeframe.
A different approach avoids the limitations of
choosing a single metric to compare emissions of
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BC and CO2, and instead investigates how reductions
of BC over the entire century would change the
difficulty of meeting radiative forcing targets.
Kopp and Mauzerall (2010) calculated the optimal
CO2 emissions pathways in order to meet a 2.21
W m 2 target in 2100. Rather than assessing the
benefits of BC reductions in the near future like the
previous studies, this study assessed the radiative
benefits of BC reductions at the end of the century
and then translated those benefits into near term
CO2 emissions targets. This study included both
co-emissions and an estimate of indirect effects.
They found that meeting this target required 50%
reductions of CO2 by about 2050. However, if this
target were tightened to accommodate the positive
radiative forcing from carbonaceous aerosols (both
OC and BC) from contained combustion source
(fossil fuels and biofuels), then the 50% reduction
of CO2 would need to occur 1 to 15 years earlier,
depending on the assumptions about carbonaceous
aerosol emissions pathways and forcing strength.
6.4.3 Valuing the Climate Benefits of BC
Mitigation
Another way to evaluate the benefits of BC
mitigation strategies and to compare them with
the benefits of other climate mitigation strategies
is to use valuation techniques to create monetary
estimates of avoided damages. This would be
equivalent to the approach adopted to compare
the health benefits of different regulatory
approaches discussed above in section 6.3. However,
methods for establishing the economic value of
the climate damages associated with BC are still
being developed. Two metrics, the Global Damage
Potential (GDP) and the social cost of a pollutant,
involve monetization of the damages of climate
change. Assessing the value of damages through a
single metric (i.e., dollars) provides useful information
that can help inform policymakers regarding the
scale and scope of the climate impacts of BC and
the benefits that can be gained from BC mitigation.
However, no study to date has fully monetized the
climate impacts of BC. An analysis of this type would
need to include the benefit of avoiding risks and
impacts associated with warming (especially near
term warming and rate of change),as well as the
value of avoiding impacts such as accelerated ice
and snow melt and changes in precipitation induced
byBC.
Currently, efforts to develop valuation methods
for climate impacts have focused on CO2. In
computing the value to society of avoided climate
damages, EPA assigns a benefits dollar value to
CO2 emission reductions using estimates of a
"social cost of carbon" (SCC) developed by a U.S.
federal government interagency working group
in 2010. The SCC is an estimate of the monetized
damages resulting from an incremental increase
in CO2 emissions in a given year; likewise, it can be
thought of as the monetized benefit to society of
reducing one ton of CO2. The SCC estimates are
intended to include an array of human-induced
climate change impacts, such as changes in net
agricultural productivity, human health, property
damages from increased flood risk, and the value
of ecosystem services due to climate change.
Current SCC values, such as those utilized by EPA to
analyze the benefits of the 2010 Light-Duty Vehicle
Greenhouse Gas Emission Standards and Corporate
Average Fuel Economy Standards (U.S. EPA, 2010c),
are subject to a number of limitations, including
the incomplete way in which the underlying climate
models capture catastrophic and non-catastrophic
impacts, the incomplete treatment of adaptation
and technological change, uncertainty in the
extrapolation of damages to high temperatures,
and assumptions regarding risk aversion. The
SCC estimates developed for CO2 have been
controversial due to the difficulty of estimating
economic impacts across nearly every sector of
the economy as well as valuation issues regarding
impacts on natural ecosystems. Furthermore, these
estimates were developed exclusively for CO2 and
are not directly transferrable to other GHGs or BC.8
It might be possible to use a similar approach to
develop a social cost specific to BC using integrated
assessment models (lAMs) that combine economic
growth, climate processes, and feedbacks between
the global economy and climate into a single
framework to translate BC emissions into economic
8 One approach that might appear tempting is to use existing
estimates for the SCC for CO2, and translate them into a social
cost for BC using metrics such as the 100-year global warming
potential, or GWP (see, for example, Copenhagen Consensus
Center Reports). However, the damage functions used in the
underlying models are sensitive to when and by how much the
temperature changes - therefore, given the orders of magnitude
shorter lifetime, a social cost calculated from first principles for
BC could be very different than one that merely scales the social
cost of CO2 by the GWP. Again, regional dependence and impacts
on precipitation patterns would not be captured by this method,
nor would the regional dependence of snow and ice deposition
and therefore special sensitivity of alpine and Arctic ecosystems
to BC emissions. Therefore, the social cost of BC might not be well
represented by using GWPs to scale an SCC. (See further discussion
of the applicability of GWP metrics to BC in Chapter 2.) Given that
warming profiles and impacts other than temperature change vary
across climate forcers, the interagency SCC working group made
a preliminary conclusion that transforming other climate forcers
"into CO2-equivalents using global warming potential, and then
multiplying the carbon-equivalents by the SCC, would not result
in accurate estimates of the social costs" of these non-CO2 forcers
(Interagency SCC Group, 2010), though it is unclear how large such
an error would be.
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Benefits of Reducing Black Carbon Emissions
(a)
CO, Emissions
Atmospheric
Concentrations
CO? Fertilization
Temperature
Change Effects
(Direct)
\
CO2 Damages
(b)
BC Emissions
Atmospheric
Concentrations
Temperature
> Change Effects
(Direct& Indirect)
Hydrological
Pattern Shift
Figure 6-9. Cause and Effect Chains for (a) CO2 and (b) BC from Emissions to Damages. (Source: U.S. EPA)
damages. However, there are a number of factors
that would complicate these calculations. These
difficulties stem partly from limitations in the
capabilities of lAMs, and partly from the complexity
of the cause-effect chain needed to measure the
physical links between emissions and climate change
impacts, and to calculate damages (see Figure 6-9).
Few lAMs are designed to demonstrate regional
impacts, and currently these models do not
adequately consider the impact of BC and other
short-lived climate forcers on the rate of climate
change. In addition, the feasibility of considering
indirect climate effects such as the impact of BC on
snow and glacier melt and changes in precipitation
patterns in the lAMs must be evaluated. In one
aspect, at least, calculating a social cost for BC might
actually be easier than calculating a social cost
for CO2: the short lifetime of BC and the relatively
immediate nature of the climate impacts reduce
the extent to which social cost calculations would
depend on the social discount rate selected. Due to
the complexities involved with valuing the climate
benefits of BC reductions, a top-down approach
using lAMs may not be preferred. Rather, a bottom-
up approach that considers location, emission
profiles of sources, and ambient concentrations and
deposition of BC similar to the approaches used to
quantify health effects may be needed.
The cause-effect chain from emissions to impacts
and damages is also complex for BC. The regional
nature of many BC impacts, the importance
of location of emissions, and BC's impacts on
precipitation, snow/ice, and surface dimming add
additional complexities to any such approach that
are not present for CO2 SCC calculations. In addition,
the peer reviewed literature lacks impact functions
and valuation methods necessary to assess many
of these BC effects. Finally, because BC is emitted
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as part of a mixture, incorporation of the climate
impacts of reducing other co-emitted aerosols into
a social cost approach would reflect the net impact
more accurately.
The Global Damage Potential (GDP) compares the
relative damage resulting from an equal mass of
emissions of two climate forcers (IPCC, 2009) -
effectively, the ratio of the social cost of a climate
forcer to the social cost of carbon. The GDP is
meant to be a parallel approach to the GWP, and
therefore might potentially be calculated using
some of the same assumptions that go into the
GWP calculation. One key assumption is that the
background concentration is a constant, modified
only by the initial pulse of emissions. This simplifies
the calculation of the GWP or GDP because
it requires no assumptions about a reference
emissions pathway. However, because the impact of
increased concentrations depends on the starting
concentration, this simplification means that the
metric may not accurately reflect actual damages.
As this discussion of GDP and SCC illustrates,
calculating economic damages associated with
specific climate forcers is extremely complicated.
Even where risks and impacts can be identified and/
or quantified with physical metrics, it may be difficult
to monetize these risks and impacts (e.g., such as
ecosystem damage or the potential to increase
the probability of an extreme weather event) such
that an accurate cost-benefit comparison could
be undertaken. Both the GDP and the social cost
calculation depend on the physical aspects of the
climate system as well as the economic linkages
between climate change impacts and the economy
(IPCC, 2009). Therefore, the GDP and the social cost
require calculations of the entire cause and effect
chain, but as a result contain a large amount of
uncertainty. Additional work is needed to design
approaches to valuing climate impacts of BC directly,
and to incorporate those approaches into metrics
comparable to the SCC.
Some authors have attempted to incorporate
economic valuation approaches into a comparative
framework that enables direct comparisons between
the benefits of BC mitigation and the benefits of
CO2 mitigation. If fully developed, such approaches
could be utilized to help policymakers choose
among an array of mitigation choices involving
different pollutants, different sources, and different
timeframes. Using a computer model that included
economic considerations, Manne and Richels (2001)
examined relative tradeoffs between different gases
that vary over time and are calculated to optimally
achieve a given target. The paper demonstrated that
if a long-term temperature stabilization target is
the only policy goal, then reductions of short-lived
gases have little value compared to long-lived gases
as long as the target will not be reached for several
decades, but that the value of these short-lived
gases rises rapidly as the temperature approaches
the target. Manne and Richels also examined a case
in which the rate of change of temperature was a
goal along with the long-term temperature change,
finding that in that case the relative prices of the
different gases stay more constant over time. This
kind of approach, including economic considerations
for cost of control but without looking at the
benefits of those controls, is known as a cost-
effectiveness analysis. The relative tradeoff between
a given gas and CO2 is also known as the Global
Cost Potential (GCP).
6.4.4 Conclusions Regarding Climate
Benefits
The climate benefits of BC mitigation are less
well understood and less certain than the health
benefits. Studies examining across-the-board
emissions reductions from individual sectors find
that the warming impact of source sectors generally
corresponds with the OC/BC ratio of that sector,
with the benefits from source sectors that have
low OC/BC ratios (e.g., fossil fuel-based sectors)
being higher.9 Most studies are consistent in
finding that the transportation sector contributes
the most to positive radiative forcing, followed by
household biofuels. Industrial sources are often
found to have a net cooling impact; however,
this broad categorization neglects to highlight
several sub-sector sources that are major BC
emitters, including brick kilns and coke ovens in
developing countries. The magnitude of the results
from studies estimating the climate benefits of
emissions reductions are often uncertain due to
uncertainty regarding co-emitted species, indirect
effects, and effects on precipitation. However, the
studies described above demonstrate that several
specific and presently available emissions control
measures are likely to have substantial climate
benefits. These include emissions control measures
for vehicles, residential burning of solid fuel, and
major industrial sources of BC including brick kilns
and coke ovens. The available literature also strongly
suggests that BC mitigation can provide particular
benefits to sensitive regions, including the Arctic
and the Himalayas. These regions stand to benefit
disproportionately from reductions in BC, especially
if reductions can be achieved from sources within
the regions themselves.
9 Bond et al. (2011) found that for direct forcing only, a ratio of
about 15:1 for OC to BC is close to climate neutral.
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Benefits of Reducing Black Carbon Emissions
Comparing the climate benefits of BC and CO2
mitigation is complicated by the many differences in
lifetime and mechanisms of impact between these
climate forcers. BC reductions can reduce the rate of
climate change and provide climate benefits in the
near term, but cannot substitute for CO2 reductions
for purposes of alleviating long-term warming. Thus,
controlling both short-lived forcers and long-lived
greenhouse gases is necessary to achieve the target
of constraining temperature rise to no more than 2°C
as agreed upon by the international community.
Several metrics have been suggested to value the
climate benefits of BC mitigation in economic terms.
Assessing the value of damages through a single
metric (i.e., dollars) provides useful information
that can help inform policymakers regarding the
scale and scope of the climate impacts of BC and
the benefits that can be gained from BC mitigation.
However, no study to date has fully monetized the
climate impacts of BC, and a great deal of additional
work is needed to design approaches for doing so.
6.5 Environmental Benefits of BC
Reductions
In addition to health and climate benefits, there
are additional environmental benefits related to
reductions in PM2.5 including BC. While EPA has
had some success in quantifying and valuing
benefits from improved visibility, other important
impacts such as ecosystem effects and damage
to building materials are not easily quantified. In
general, however, the environmental benefits of
reducing PM2.5 have been shown to be quite large
in the United States. Importantly, the majority of
environmental benefits globally are likely to accrue
to other countries, i.e., those investing in emission
reduction programs. This is particularly true for those
areas where ambient PM25 is interfering with rainfall
patterns (discussed above, in the section on climate
impacts) or causing surface dimming on a broad
scale. The next sections describe what is known
about how PM2.5 reductions can (1) improve visibility,
(2) reduce pollutant impacts on ecosystems, and (3)
enhance the longevity of building materials.
6.5.1 Visibility Impacts
Visibility impairment is caused by the scattering and
absorption of light by suspended particles and gases
in the atmosphere. A number of other factors can
influence visibility, such as the relative atmospheric
humidity, intensity of sunlight, presence of cloud
cover, distance from the object being viewed,
physical characteristics of the object being viewed,
and physical capabilities (i.e., eyesight) of the viewer
(Malm, 1999). However, when PM2.5 is present in
the air, its contribution to visibility impairment
typically greatly exceeds that of naturally occurring
atmospheric gases (U.S. EPA, 2011d). As a result, in
otherwise constant conditions, visibility impairment
is greater when PM is present. Reductions in
air pollution from implementation of various
programs associated with the Clean Air Act (CAA)
Amendments of 1990 provisions have resulted
in substantial improvements in visibility, and will
continue to do so in the future.
Visibility directly affects people's enjoyment in a
variety of daily activities and their overall sense
of well-being. Individuals value visibility both in
the places they live and work, in the places they
travel to for recreational purposes, and at sites
of unique public value, such as national parks.
Visibility economic benefits consist of the aesthetic
benefits of better visibility, improved road and air
safety, and enhanced recreational activities like
hunting and bird watching. As with health benefits,
visibility improvements are valued using WTP
studies. EPA estimates that in 2010, improvements
in visibility related to the 1990 CAA Amendments
were valued at approximately $36 billion annually
(2010$) (U.S. EPA, 2011c). Almost three-quarters of
these benefits ($27 million) result from residential
visibility improvements. It is important to note that
residential benefits in this EPA study reflect benefits
from all metropolitan statistical areas in the country,
whereas recreational benefits are limited to visibility
improvements in Class I areas managed by the
National Park Service in California, the Southeast
and Southwest.
6.5.2 Ecosystem Impacts
Ecosystems perform a number of functions that
contribute to human welfare, including the provision
of food and raw materials, filtering of air and water,
and protection from natural hazards such as floods
(U.S. EPA, 2011d). Additionally, people may seek
out certain ecosystems for their aesthetic value.
Atmospheric PM2.5 negatively affects the ability of
ecosystems to perform these and other valuable
welfare functions. PM25 can impact ecosystems
through direct deposition on plants, animals, or
bodies of water. In areas with high emissions, PM2.5
deposition on leaves interferes with a plant's ability
to perform basic metabolic functions. Increases in
trace metal and organic matter in bodies of water
as the result of PM25 deposition are toxic for aquatic
life forms (U.S. EPA, 2009b). Indirect impacts are
caused when plants and animals take in pollutants
through affected soil or water. As with direct
impacts, indirect impacts can alter normal biological
processes and be toxic to living organisms.
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The relationship between PM2.5 and ecosystem
effects is difficult to quantify because of the
variability in PM2.5 emissions and composition.
Pollutants accumulate over time in affected
organisms, making it difficult to link impacts to
ambient PM2.5 concentrations. Additionally, the
impact of negative welfare effects may vary based
with geographic location. The impact of damage to
a national park would likely be valued differently
than damage to commercial farm land, which would
in turn be valued differently than damage to private
non-commercial property (U.S. EPA, 2011d).
6.5.3 Materials Co-benefits
PM2.5 deposition on materials such as stone,
metal, and painted surfaces leads to damage
by accelerating the natural weathering process.
Chemical reactions with acidic gases worsen the
impact of PM2.5 -related damage. Additionally,
accumulation of PM25 on surfaces, referred to as
soiling, affects the aesthetic properties of materials
and necessitates more frequent cleaning or
repainting of affected surfaces. Research has not
established any quantitative relationship between
the ambient concentrations of PM2.5 and the rate of
damage or soiling caused by PM25 deposition (U.S.
EPA, 2011d).
6.5.4 Conclusions Regarding Environmental
Benefits
The environmental benefits of reducing BC are likely
to be substantial, both in terms of the range of
impacts avoided and the value to society, although
it is difficult to quantify these impacts currently. Due
to the difficulties involved in quantifying and valuing
environmental benefits, EPA often addresses these
benefits qualitatively with the exception of visibility
benefits as previously discussed.
6.6 Conclusions
All control measures that reduce PM2.5 pollution
are virtually certain to achieve health benefits,
and several studies examining costs and benefits
of BC mitigation suggest that the health benefits
alone may justify mitigation. Programs to reduce
fine particles in the United States and in other
developed countries have greatly reduced the
negative health impacts of PM25, including BC. EPA
has determined that there is insufficient information
at present to differentiate the health effects of the
various constituents of PM2.5; thus, EPA assumes
that many constituents are associated with adverse
health impacts. New programs introduced by EPA
for mobile and stationary sources will continue to
reduce PM2.5-related health impacts in the United
States over the next several decades. The largest
potential benefits of BC mitigation measures are
achievable internationally, due to high emissions co-
located with large populations, particularly in South
and East Asia. More information on the benefits
and costs of individual measure in each country is
needed to support policy decisions made at the
national level.
Estimating the climate benefits of BC mitigation
is less well understood and less certain compared
with estimating health benefits. However, several
conclusions can be drawn from the literature
examining the climate impacts of BC reductions.
Current studies indicate that BC reductions can
reduce the rate of climate change in the near-
term. Controlling emissions from motor vehicles
and residential burning of solid fuels is likely to
benefit climate, though residential emissions are
particularly difficult to estimate and errors in current
understanding of the composition of emissions
may affect this conclusion. Major industrial
sources of BC, such as brick kilns and coke ovens,
will likely also lead to climate benefits. There are
several key uncertainties which further research is
needed to address, including the extent to which
different emissions mixtures result in equivalent
climate effects, how the indirect effects of those
mixtures influence the climate outcomes, and the
potential benefit of various mitigation strategies
for precipitation and meteorology. Additional
work is needed to design approaches to valuing
climate impacts of BC directly, and to incorporate
those approaches into useful metrics for evaluating
policy decisions, similar to the social cost of
carbon (SCC). It is also important to note that BC
mitigation cannot substitute for CO2 reductions for
the purposes of alleviating long-term warming. The
literature suggests that mitigating both short-lived
climate forcers and long-lived greenhouse gases is
necessary to achieve internationally agreed upon
goals of temperature rise.
Overall, the literature points to substantial health
and climate benefits of BC mitigation from
some sources, particularly for control measures
targeting emissions from motor vehicles, residential
combustion of solid fuels, and some high BC-
emitting industrial sources such as brick kilns and
coke ovens. Mitigation measures for each of these
sectors exist and have been proven to be effective
in different parts of the world, including in the
United States, as detailed in Chapters 8-11. Chapter
7 describes how the information presented in this
chapter can be used to evaluate policy options.
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Chapter 7
Mitigation Overview: Designing
Strategies for Public Health and
Near-Term Climate Protection
7.1 Summary of Key Messages
• Existing particle control programs have been
effective in reducing BC in many regions,
particularly control programs affecting emissions
from mobile and stationary sources.
- While BC is not the direct target of existing
programs, it has been reduced through
controls aimed at reducing ambient PM2.5
concentrations and/or direct particle
emissions.
- Past experience suggests that available
control technologies and approaches can
reduce BC emissions from many key source
categories at reasonable cost. However,
information is currently limited regarding the
effectiveness of control strategies for reducing
BC in a targeted fashion and the associated
costs of those strategies.
• While global BC emissions are likely to decrease
in the future, this trend will be dominated by
emissions reductions in developed countries and
may be overshadowed by emissions growth in key
sectors (transportation, residential) in developing
countries, depending on growth patterns.
- Developed nations have already made
significant progress in reducing BC emissions,
and further reductions are expected to
occur through 2030 with full implementation
of existing regulations particularly in the
transportation sector.
- Emissions projections for developing countries
are more variable, with studies indicating
that emissions are likely to increase in some
sectors and regions and decrease in others.
• Available control technologies can provide low-
cost reductions in BC emissions from a number
of key source categories. BC emissions reductions
are generally achieved by applying technologies
and strategies to improve combustion and/or
control direct PM2.5 emissions from sources.
- Some of the strategies utilized by developed
countries have also been undertaken in
developing countries or could be adopted on
a broader scale internationally. In other cases,
developing countries have a different mix
of sources and constraints that may require
different types of control strategies.
In selecting BC mitigation measures, policymakers
must consider three overlapping goals: climate
benefits, health benefits, and environmental
benefits. In most cases, policymakers will seek
to achieve multiple goals simultaneously; this
requires taking into account a suite of impacts
and attempting to maximize co-benefits and
minimize tradeoffs across all objectives (health,
climate, and environment).
With a defined set of goals, policymakers can
evaluate the "mitigation potential" within each
country or region. The mitigation potential
depends on total BC emissions and key emitting
sectors, and also depends on the availability of
control technologies or alternative mitigation
strategies.
Selection of ideal emissions reduction strategies
will depend on a range of constraining factors,
including:
- Timing
- Location
- Atmospheric Transport
- Co-emitted Pollutants
- Cost
- Existing Regulatory Programs
- Implementation Barriers
- Uncertainty
Considering the location and timing of emissions
and accounting for co-emissions will improve
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the likelihood that mitigation strategies will be
beneficial for both climate and public health.
- While all PM mitigation strategies that
reduce human exposures will benefit health,
strategies that focus on sources known
to emit large amounts of BC—especially
those with a high ratio of BC to OC, like
diesel emissions—will also maximize climate
benefits.
- The location and timing of the reductions
are also very important. The largest climate
benefits of BC-focused control strategies may
come from reducing emissions affecting the
Arctic, Himalayas and other ice and snow-
covered regions. This would include BC
emitted directly in those areas as well as BC
transported into those areas from other areas.
Cost is a prime consideration in both developed
and developing countries, as is feasibility of
implementation. Some physical and political
constraints may hinder full implementation of
even those strategies for which there is high
confidence of large health and climate benefits.
Optimizing climate, public health and
environmental benefits requires a broad,
multi-pollutant approach to BC mitigation that
includes looking at the entire suite of options and
evaluating them carefully to understand the full
range of costs and benefits.
7.2 Introduction
As outlined in the previous chapter, reducing BC
emissions has tremendous potential for improving
global public health while achieving climate benefits.
The optimal path forward will vary as decision-
makers in each country weigh desired health and
environmental outcomes, costs and benefits, and
mitigation potential. This is a complex calculus that
depends on a large number of considerations. This
chapter presents a decision framework to help guide
policymakers who want to develop BC mitigation
strategies.
First, the chapter examines what is known about
the overall impact of existing or planned control
programs on emissions of BC and how current
BC emissions are projected to change over
the next several decades in response to these
control programs and/or economic growth and
development. Next, the chapter describes a decision
framework, which includes both key factors to
consider when developing a BC mitigation strategy
and how different approaches for reducing BC
emissions have potential to provide climate and
public health benefits. The chapter concludes with
several examples of how the decision framework
could help guide mitigation choices. The chapter
is followed by more detailed mitigation chapters
covering four major emissions sectors—mobile
sources; stationary sources (including both power
generation and industry); residential heating and
cooking; and open biomass burning. Chapters 8-11
describe projected changes in emissions in these
sectors in the United States and globally, available
control technologies and strategies and their
associated costs, and implementation challenges.
Chapter 12 provides a summary of how the
mitigation options in these various sectors translate
into near-term mitigation opportunities for BC to
benefit climate and public health.
7.3 Effect of Existing Control
Programs
Many existing control programs have been highly
effective in reducing BC, however, it is important to
note that BC is not the direct target of any currently
existing emissions control program. Rather, BC has
been reduced through control programs focused
on reducing ambient PM2.5 concentrations or
direct particle emissions in general. As discussed
throughout this report, BC is always co-emitted with
other particles and gases. Therefore, determining
the effect of various mitigation strategies on BC
emissions requires an understanding of the entire
emissions mixture coming from a given source and
the extent to which the BC fraction is reduced by
specific control technologies or strategies. Currently,
there is only limited information about effective
control strategies for reducing BC in a targeted
fashion and the associated costs of those strategies.
In recent years, the overarching PM2.5 control
program for stationary sources in the United States
and Europe has focused mainly on secondarily
formed particles such as sulfates and nitrates,
rather than on direct PM2.5 emissions. This is
because PM controls motivated by public health
and environmental goals are focused on reducing
total PM mass (a large portion of which is sulfates
and nitrates formed in the ambient air from SOx
and NOJ at least cost. PM controls oriented toward
climate would have to consider the light absorbing
and scattering properties of the various PM
constituents.
Controls on direct PM2.5 emissions do affect
emissions of BC and other constituents such as
OC. This is clear from the limited emissions testing
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data and the observational record that link declining
BC concentrations to PM2.5 control programs (see
Chapters 4 and 5). However, the extent to which BC
has been controlled as a component of an overall
PM2.5 mixture has depended somewhat arbitrarily
on the proportion of BC in the emissions mix from
a particular source category and the specific control
strategy applied. Some strategies in some sectors
(such as mobile source emissions standards for
diesels) effectively reduce BC emissions as much
or more than other constituents, while in other
instances, BC reductions may be proportionally
smaller. In addition to uncertainty regarding the
composition of PM emissions from many sources,
the relative effectiveness of a particular control
technology for reducing specific constituents is
often unknown, which means that for most sectors,
it is not clear whether PM2.5 controls will reduce
BC preferentially or even proportionally to other
constituents. Ongoing research will help to clarify
this issue.
In general, available estimates of BC emissions
reductions are calculated from analyses of PM2.5
controls. As discussed in Chapter 4, EPA's trends
report (20101) shows that U.S. emissions of direct
PM25 have declined by 58% since 1990, a reduction
of over 1.3 million tons. Over half of this reduction
has come from controls on stationary fossil fuel
combustion, with substantial reductions also
occurring in emissions from industrial processes
and mobile sources. Using speciation factors, it is
possible to calculate BC reductions in these sectors,
but these estimates are generally rough. Information
on BC reductions is strongest for the mobile source
sector, where the BC fraction of emissions and the
impact of specific controls are well understood.
As discussed in more detail in Chapter 8, mobile
source BC emissions declined by approximately 32%
between 1990 and 2005. For other sectors, precise,
measured data about the effectiveness of specific
controls for reducing BC emissions is often not
available. As described in Chapter 5, however, recent
ambient BC measurements do appear to indicate a
decline in neighborhood/urban and regional scale
concentrations of BC in the United States between
the mid 1980s and the present (see section 5.4.1).
While control strategy information and cost data
for BC mitigation approaches are generally limited,
this varies by sector and location. Some of the best
information is available for mobile source controls.
Analyses conducted for recent regulatory actions
in the United States provide a solid foundation for
understanding applicable technologies and costs,
and related implementation issues. For other sectors
where less information is available, for example open
biomass burning, better information on BC-specific
control strategies, effectiveness and costs is needed.
EPA has historically evaluated PM control strategies
for specific sectors as part of the regulatory impact
analyses for specific rulemakings. These analyses
generally include best-available information on
control options, effectiveness, and costs. Some of
them include information on controls for specific PM
constituents, but this rarely includes BC.
Despite what is known from analysis conducted in
the United States, many of the strategies that have
been applied domestically differ in important ways
from control strategies that have been adopted
internationally. Some of the strategies utilized by
developed countries have also been undertaken
in developing countries or could be adopted on
a broader scale internationally. In other cases,
developing countries have a different mix of sources
and other relevant constraints that will require
different types of control strategies. These issues are
discussed further in the sections that follow, and in
the conclusion to this chapter.
7.4 Future Black Carbon Emissions
The influence of BC on climate and public health
in the future, and the need to more precisely
determine the effectiveness of various mitigation
strategies for reducing BC, depends in large part
on the magnitude of future emissions. This section
describes what is known regarding these future
emissions, but available estimates are variable and
uncertainty about future emissions trajectories
remains high.
Developed nations have already made significant
progress in reducing direct PM emissions,
and further reductions are expected to occur
through 2030 with full implementation of existing
regulations. In particular, substantial BC reductions
have been achieved through controls in the mobile
source sector (particularly diesels), and additional
reductions will continue to be realized over the next
two decades. In the case of stationary sources, the
most substantial BC emissions reductions in the
United States and other developed countries were
achieved decades ago (often through fuel switching
away from coal).
Recent studies (Streets et al., 2004; Cofala et al.,
2007; Jacobson and Streets, 2009; Rypdal et al.,
2009) provide a snapshot of potential future BC
emissions trends. These studies have produced
a range of estimates for future BC emissions
depending on assumptions about economic growth,
population levels, and development pathways. In
an analysis of future BC emissions trends based
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Chapter 7
o
01
c
g
8 2
LL
(a)
A1B
CD Residential Q Power
Q Industrial fj Transport
A2
B2
B1 F A1B
(b)
Savanna Q Ag Res
Trop Forest Q Ex-Trop Forest
A1B
A 113
A2 B2
B1
A2 B1
1996
2030
2050
1996
2030
2050
Energy-related combustion
Open biomass burning
Figure 7-1. Global BC Emissions Forecasts for Various Sectors under Alternative IPCC SRES Scenarios (in
teragrams (Tg) of carbon). Scenarios generally show a modest decrease in BC emissions from all sectors as
compared to 1996 baseline emissions. (Streets et al., 2004)
on the IPCC SRES scenarios, Streets et al. (2004)
projected BC emissions to decrease globally by 9%
to 34% by 2030 relative to 1996 levels depending
on assumptions about economic growth and
development. However, there was considerable
variation among projections for the different sectors
depending on the SRES scenario examined (Figure
7-1). Thus, while aggregate emissions were generally
projected to decline under alternative growth
scenarios, emissions growth was projected for
certain sectors or regions. The sectors where Streets
et al. (2004) indicate a potential for future emissions
growth include residential emissions in Africa, open
biomass burning emissions in South America, and
transportation emissions in the developing world
(for example, where fuel sulfur levels are still too
high for implementation of DPFs—see Chapter 8
and Appendix 3). In general, industrial emissions
and transport emissions were projected to decline in
developed countries.
An analysis byJacobson and Streets (2009) found
that under the assumptions embedded in the A1B
scenario for IPCC, total global BC emissions may
increase substantially. Again however, this analysis
indicates that projected emissions growth or decline
varies significantly among regions and sectors,
as Figure 7-2 illustrates. In general, BC emissions
in developed countries are projected to decline,
while emissions in developing countries may grow.
Transportation (mobile source) emissions in particular
are projected to grow in several world regions but
decline in others, as illustrated by growth factors
greater than or less than one, respectively (Jacobson
and Streets, 2009). In developed countries, the
majority of the emissions reductions in the
transportation sector are projected to result from
implementation of 2007 U.S. on-highway diesel
engine standards and similar standards, such as
Euro V, that lead to the use of diesel particulate
filters (DPFs) in the diesel fleet. Also, in the United
States, other standards for diesels (nonroad diesels,
locomotives, and commercial marine) contribute
to these reductions. However, other studies have
indicated that emissions from shipping in the Arctic
region may increase due to the retreat of Arctic sea
ice, opening up new shipping routes and increased
economic activity in that region (Corbett et al.,
2010).
In its most recent work, the IPCC has also developed
four "Representative Concentration Pathways"
for use as a consistent set of emissions inputs
for projecting future climate change. These four
pathways (Figure 7-3) are defined by the total
radiative forcing resulting from each pathway in
2100, including GHGs and other forcing agents,
which ranges from 2.6 to 8.5 W m2. Global BC
emissions in all four pathways peak in 2005 or
2010, are 8 to 20% below 2010 levels by 2030, and
continue decreasing for the rest of the century
to about half of 2010 levels. Emissions for the
764
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Mitigation Overview: Designing Strategies for Public Health and Near-Term Climate Protection
(a)
3 n
2
(D)
1 4
f
0
n
•
mm
r
n
D(
• E
Carbon monoxide
Slack carbon
...
-
E
n
r
-
11 rl
M
Residential FF Industry Transportation
Biomass
Burning
o^^VVV^ /^"^ /" V^ V
Figure 7-2. Black Carbon Emissions Growth, 2000-2030 under IPCC A1B Scenario. Top: 2000-2030 Black
Carbon Emissions Growth Factors by Sector for Selected World Regions (from IPCC A1B scenario). Bottom:
2000-2030 Black Carbon and Carbon Monoxide Emissions Growth Factors for Transportation (Mobile)
Sector in Specific Regions. Emissions in sectors with a growth factor less than one (see dark line, added)
will decline. (Source: Jacobson and Streets, 2009)
8-
7-
? 6~
£ 5-
w
I 4
'£ Q
LJJ 3
2-
1 -
J I
Scenarios
— AIM-RCP6.0
IMAGE - RCP3-PD (2.6)
— MiniCAM - RCP 4.5
MESSAGE - RCP 8.5
2000
2020
2040
2060
2080
2100
Year
Figure 7-3. Future Emissions of BC under IPCC Representative Concentration Pathways, 2000-2050
(Gg/year).
Notes:
1. RCP2.6 (RCP 3-PD) - van Vuuren et al., (2007)
2. RCP 4.5 - Clarke et al. (2007); Smith and Wigley (2006); and Wise et al. (2009)
3. RCP 6.0 - Fujino et al. (2006); and Hijioka et al. (2008)
4. RCP 8.5 -Riahietal. (2007)
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Chapter 7
RCP pathways are reported in combinations of
five regions and four sectors on the website1
(though the underlying, gridded dataset is further
disaggregated).
Consistent with the findings of the academic studies,
under certain pathways there are a few region and
sector combinations whose BC emissions do not
peak until 2020 or 2030. Two out of the four RCP
pathways show near-term increases in BC emissions
from all sectors in Asia, the Middle East, and Africa,
and for all regions some of the pathways show
increases in open burning emissions (from both
deforestation and agricultural burning). Because
of the potential for increases in open burning, OC
emissions are not projected to decline as quickly as
BC emissions.
BC emissions in the United States are projected
to decline, driven largely by reductions in mobile
diesel emissions, as discussed in detail in Chapter 8.
The limited EPA modeling inventories that project
emissions into the future (year 2020) indicate that
direct PM2.5 emissions from industrial sources are not
expected to decline significantly in the next decade,
and emissions from fossil fuel combustion will
only decline about 20% by 2020 (U.S. EPA, 2006c).
Because of the small size of anticipated reductions
in direct PM2.5 emissions from these categories,
projected BC emissions changes are also small
and unlikely to affect the U.S. BC emissions trend
in the future in the absence of additional control
requirements. Open biomass burning, the second
largest source category in the United States, exhibits
significant year-to-year variability in emissions, and it
is difficult to predict future year emissions. However,
it should be noted that emissions in this category
may grow significantly in the future if climate change
results in increased wildfires, as predicted in many
scenarios (Wiedinmyer and Hurteau, 2010).
Projected future emissions reductions may not occur
in the United States or elsewhere in the absence of
continued policies to encourage adoption of DPFs in
the mobile sector, continued economic development
leading to a more rapid shift away from traditional
cookstoves than is currently predicted, and other
environmental and economic developments. As
noted, there are also several sectors and regions,
such as transport emissions in developing nations
and open biomass burning emissions globally,
for which emissions are not projected to peak for
another decade or two. Given the array of available
control technologies and strategies, as outlined in
the next several chapters of this report, it is possible
to make larger and more rapid reductions in BC
emissions globally than current baseline estimates
project.
Some countries have already begun looking at
these possibilities. For example, the Arctic Council
countries (Canada, Denmark, Finland, Iceland,
Norway, Russia, Sweden, and the United States)
formed a special Task Force on Short-Lived Climate
Forcers in 2009 to consider whether additional or
accelerated mitigation strategies may be needed
to address warming in the Arctic region. Noting
that emissions in the Arctic region from sources
other than land-based transport—particularly
residential heating, agricultural and forest burning,
and marine shipping—will likely remain the same
or increase without new measures, the Task Force
has recommended that "Arctic Council nations
individually and collectively work to implement
some early actions to reduce black carbon" (Arctic
Council Task Force on Short-Lived Climate Forcers,
2011, p. 5). A new Project Steering Group (PSG)
under the Arctic Council's Arctic Contaminants
Action Programme (ACAP) is currently investigating
which measures would be implemented, and the
impact of such measures on total BC emissions
from Arctic Council nations. The PSG has identified
a number of key sources affecting the Arctic and is
developing pilot programs to advance BC emissions
reduction efforts, particularly in the Russian far
north.
7.5 Key Factors to Consider in
Pursuing BC Emissions Reductions
Significant reductions in BC emissions are expected
to occur in certain regions in the coming decades;
however, these reductions will be gradual, and even
after they are fully realized, substantial BC emissions
will remain in some sectors and regions. There is a
core set of factors that policymakers can consider
to improve the likelihood that selected mitigation
strategies will achieve substantial public health
and environmental benefits and reduce the rate of
near-term warming. Policymakers can examine the
challenge of BC mitigation from the perspective
of three different (but overlapping) goals: climate
benefits, health benefits, and environmental
benefits. With a defined set of goals, policymakers
can evaluate the "mitigation potential" within each
country or region. The mitigation potential depends
on total BC emissions and key emitting sectors,
and also depends on the availability of control
technologies or alternative mitigation strategies,
such as fuel switching, improvements in energy
efficiency, or changes in land-use patterns. The ideal
emissions reduction strategies will also depend on a
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Mitigation Overview: Designing Strategies for Public Health and Near-Term Climate Protection
JION POTENTIAL +/- CONSTRAINING FACTORS
Goals
Climate
Radiative Forcing
Temperature
Ice/Snow Melt
Precipitation
Health
Ambient Exposures
Indoor Exposures
Environment
Surface Dimming
Visibility
Emissions sources
Stationary
Sources
Brick Kilns
Coke Ovens
Diesel Generators
Utilities
Flaring
Mobile
Sources
On-Road Diesel
On-Road Gasoline
Construction Equip.
Agricultural Equip.
Locomotives
Open Biomass
Burning
Agricultural Burning
Prescribed Burning
Wildfire
Residential
Cooking and
Heating
Cookstoves
Woodstoves
Hydronic Heaters
Mitigation options
Available Control
Technologies
e.g. Diesel
Particulate Filters
Alternative Strategies
to Reduce Emissions
e.g. Efficiency
Improvements, Substitution
Timing
Location
Atmospheric
Transport
Co-Emitted
Pollutants
Cost
Existing Regulatory
Programs
Implementation
Barriers
Uncertainty
Figure 7-4. Policy Framework for Black Carbon Mitigation Decisions. (Source: U.S. EPA)
range of constraining factors, such as cost, location,
and co-emitted pollutants. This decision-making
framework is illustrated in Figure 7-4. Each of the
key factors policymakers must consider is discussed
further below.
7.5.1 Defining Goals: Climate, Health and
Environmental Outcomes
Reducing BC emissions offers a win-win opportunity
with the potential to achieve benefits for climate,
public health and the environment simultaneously.
The preferred mitigation strategies could differ
depending on the main policy goal. Policymakers
focused primarily on climate objectives might choose
different mitigation approaches than policymakers
who are interested in maximizing public health
benefits or protecting visibility. BC plays an
important role in all of these effects, but achieving
different goals might require different mitigation
strategies oriented toward different sources in
different locations. In most cases, policymakers
will be seeking to achieve multiple goals at the
same time; thus, preferred strategies will likely
take into account a suite of impacts and decision-
makers will attempt to maximize climate, health
and environmental benefits and minimize tradeoffs
across all objectives. However, it is important for
policymakers to be clear about what the goals are,
and to prioritize among them when tradeoffs are
involved.
Often, policymakers will be considering BC
reductions as merely one part of a broader PM
mitigation agenda, and decisions will be driven first
by health considerations. This is consistent with
the mitigation pathway in developed countries,
where the primary goal has been improving public
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Chapter 7
health, and the secondary goal has been reducing
non-climate environmental effects of PM (such
as acid deposition and visibility impairment). PM
mitigation programs in most countries have not
given direct consideration to climate benefits.
Policies have generally focused on reducing
emissions of secondary PM precursors (SO2 and NOX)
as the most cost-effective strategy for those health
benefits. While enormously successful for public
health, strategies that reduce SO2 and NOX emissions
to control secondarily formed PM are not expected
to result in substantial decreases in BC emissions.
Moving forward, as climate becomes an additional
consideration, SO2 and NOX reductions should be
accompanied by corresponding reductions in BC.
The practical result would be additional policies that
more actively target sources of direct PM emissions.
This would ensure that ongoing reductions in SO2
and NOX emissions, which are critical to achieving
public health and non-climate environmental goals,
are complemented and enhanced by BC emissions
reductions for climate.
Even within a specific category (climate, health,
environment), there are multiple sub-goals to be
considered. While related, these different objectives
may require different strategies and approaches.
For example, BC reductions could provide climate
benefits in the form of reduced radiative forcing and
temperature, but could also be aimed at reducing
precipitation impacts and the melting of ice and
snow. The optimal control strategy will depend
on the impact(s) of concern. Scientific evidence
discussed in Chapter 2 suggests that direct radiative
forcing and temperature increase are driven
largely by BC emissions, offset by emissions of
other aerosols. Reducing these impacts, therefore,
may require strategies that preferentially reduce
BC relative to cooling species. On the other hand,
a broader variety of aerosols (including nitrates,
sulfates, OC) are contributing to snow/ice albedo
effects, ABCs, and precipitation effects, so to
reduce these impacts, a wider variety of PM2.5
reduction strategies may be beneficial. For health,
policymakers are interested in both indoor and
outdoor exposures and risks. However, reducing
indoor exposures to PM2 5 (including BC) often
requires different strategies than reducing ambient
PM2.5 concentrations. Identifying clearly which set of
impacts the mitigation action is designed to reduce
is critical for selecting appropriate measures.
Because of the nested nature of the climate, public
health and environmental goals driving policy
decisions about BC mitigation, cost-benefit analysis
conducted to support mitigation decisions should
incorporate public health and welfare benefits as well
as climate benefits. Such analysis can fully inform
decision makers regarding available choices, but can
also be complicated by uncertainties in valuing the
climate impacts. This is particularly true for other
environmental outcomes, such as for visibility and
impacts on agriculture. The impact of BC on these
outcomes is less well understood, and therefore
there is less known quantitatively about expected
benefits of BC reductions. Nevertheless, as discussed
in Chapter 6, public health benefits of many BC
reduction strategies may be large enough to justify
the costs, regardless of the climate impacts.
7.5.2 Identifying Opportunities for
Emissions Reductions
Designing BC mitigation programs requires
policymakers to carefully evaluate both the total BC
emissions inventory in a specific region or country,
and the contribution of specific sectors to that total.
As discussed in detail in Chapter 4 and mentioned
earlier in this chapter, BC emissions—and therefore
BC mitigation potential—vary widely by region. In
some countries, aggressive programs to reduce PM
emissions mean that the remaining emissions are
limited, making further reductions more difficult
and expensive to achieve. Nevertheless, virtually
every country, including the United States, still
has substantial BC emissions remaining across a
number of source categories, meaning that further
BC reductions are possible. For example, while U.S.
BC emissions are expected to drop considerably
by 2030 due to controls on new mobile diesel
engines (see Chapter 8), significant emissions
remain from the existing fleet of diesel vehicles.
Achieving additional reductions from mobile diesel
engines is currently possible through EPA's National
Clean Diesel Campaign and SmartWay Transport
Partnership Program. However, funding would be
required in order to address a majority of the legacy
fleet of existing vehicles. Furthermore, as discussed
in Section 7.4, several recent studies identify
regions and emissions source categories in which
BC emissions are likely to increase over time. These
sectors represent important potential mitigation
opportunities, even though total emissions from
some of them may not be large at present.
Careful investments in emissions inventories and
emissions measurements can greatly improve
policymakers' ability to identify key emitting sectors
and sources. This includes evaluating current
emissions and anticipated future BC emissions in
terms of both specific facilities (how many? where?)
and control technologies already in place. Having
a clear understanding of exactly where potential
emissions reductions could be achieved based on
current sources and technologies opens the door for
practical conversations about specific avenues and
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means for achieving those reductions. This includes
considering both:
• Availability of control technologies: As
individual countries examine the options for
reducing BC to improve air quality and benefit
the climate, they must consider the availability
of control technologies that are effective for
BC mitigation. Technologies to reduce PM and
co-pollutant BC emissions are in widespread use
globally, as are technologies that reduce PM and
co-pollutants more broadly.
• Alternative mitigation strategies: Even when
conventional control technologies are ineffective
or costly, other strategies may be available
to reduce BC emissions. In some cases, BC
reductions may be achieved not through end-of-
pipe controls (such as particle filters) but through
substitution of new, cleaner technologies (such
as improved cookstoves and brick kilns) or more
efficient combustion practices that substantially
alter the emissions stream. Changes in land-use
patterns or greater reliance on alternative forms
of energy can also result in BC reductions.
Previous chapters have covered in detail what
is known regarding emissions sources from the
United States and across the globe. The following
chapters will go into greater detail regarding specific
technologies and strategies available to control
BC emissions from the main contributing sectors.
Appropriate mitigation choices for individual
countries will vary based on scientific variables and
policy drivers.
7.5.3 Key Considerations
Even with clearly defined goals and carefully
constructed emissions inventories, there are a
number of other factors for policymakers to consider
that are critically important for mitigation decisions.
These include:
• Timing: Seasonal timing of emissions reductions
is important. As discussed in Chapter 2, the
effects of BC in snow- and ice-covered regions
are accentuated during times of increased sun
exposure. Policies designed to reduce BC impacts
on the Arctic, therefore, might need to consider
the seasonality of emissions. For example, impacts
of open biomass burning on the Arctic spring
melt could be influenced by adjusting the timing
of agricultural burning. Similarly, the benefits of
emissions controls for the Indian Monsoon will
depend on the timing of controls relatively to
seasonal weather patterns.
Location: Considering the location of emissions
is also important when seeking mitigation
strategies that will be beneficial for both climate
and public health. Because BC is a regional rather
than global pollutant, it is important to evaluate
the benefits of emissions reductions in terms
of specific regional impacts, factoring in source
region, emissions transport, and receptor region
conditions. For climate, impacts on snow- and
ice-covered regions such as the Arctic are of
particular concern, as are impacts in regions
where the precipitation patterns are heavily
impacted by all aerosol emissions, such as India.
Location matters for human health benefits too,
with proximity of emissions reductions to areas
of large population being a prime consideration.
In general, the largest human health benefits
from BC-focused control strategies occur near
emissions sources, where exposure affects a
large population. Despite the historical focus
on secondary PM for public health, many of the
BC sources ripe for mitigation are in large urban
areas. These sources of direct PM, from the
transportation, residential and industrial sectors,
are tied to everyday human activities. This means
that the benefit of reducing BC emissions for the
health of the people living nearby is expected to
be very high.
Emissions Transport: The net impacts of
emissions from any specific source will depend
partly on atmospheric fate and transport.
Meteorological conditions, plume height and
other factors will affect the extent to which
emissions from particular sources in particular
locations affect climate, health and environmental
endpoints of interest. Identifying how reductions
in emissions from specific sources translate into
climate, health and environmental responses in
receptor regions requires sophisticated models
that can account for myriad chemical and physical
processes, as well as climate responses.
Co-emitted pollutants: Maximizing benefits
across different goals requires explicitly
accounting for co-emitted pollutants. Depending
on the goal, reductions in some species may
be more valuable than reductions in others.
The largest benefits in terms of direct forcing
would come from two types of strategies: first,
strategies that reduce BC more than emissions
of other (cooling) PM constituents like SO2,
NOX and/or OC; and second, strategies that
reduce BC in conjunction with reductions in
other GHGs like CO2 and CH4. The first approach
would include strategies such as controls on
mobile diesel engines (see Chapter 8), while
the second approach would include strategies
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Chapter 7
Open Biomass Burning
U.S. Diesel Vehicles
Global On-Road Vehicles
H^^^H
Global Non-Road Vehicles
Global Industrial
U.S. Gasoline Vehicles
U.S. Fossi Fue Combustion
Global Residential
Global Agricultural Burning
Global Power Generation
Figure 7-5. OC (left) and BC (right) Emissions from Key U.S. and Global Emissions Source Categories,
Expressed as a Fraction of Total Carbon (OC + BC) Emissions from that Category. (Source: U.S. EPA)
such as improvements in combustion and
fuel efficiency in residential cookstoves, use
of prepared fuels (briquettes or pellets), or a
shift to clean cooking fuels (see Chapter 10).
For both types of strategies, optimizing control
opportunities requires careful analysis of the full
emissions stream. When all climate effects (i.e.,
effects beyond direct forcing) are considered,
and particularly when health and environmental
benefits are added to the equation, the relative
importance of various co-pollutants changes.
For reducing direct forcing, mitigation strategies
might focus on reducing BC preferentially. If the
focus shifts to health impacts or precipitation
effects, other aerosol species become equally
important and total PM2.5 reductions are likely to
be preferred over BC-focused strategies.
It is very difficult to establish short-hand approaches
for estimating the impacts of specific sources
based on their chemical emissions profile. In a
number of studies, the OC/BC ratio2 has been used
to rank the net warming potential of individual
source categories, based on an assumption that
OC will primarily reflect light and thereby induce
a negative radiative forcing (cooling) effect in the
atmosphere, and that BC will primarily absorb light
and induce a positive forcing (warming) effect.
Figure 7-5 illustrates the variation in emissions
profiles among sources by comparing relative OC/
BC ratios. Emission sources with low OC/BC ratios
are generally thought to have the largest potential
to warm the climate, though there is no agreement
within the scientific literature about how to interpret
specific ratios. Bond et al. (2011) found that for
direct forcing only, a ratio of about 15:1 for OC to
BC is close to climate neutral; however, this does
not include cloud indirect effects or co-emissions of
substances other than OC. Bond et al. also find that
the neutral ratio varies substantially depending on
where the emissions occur.
2 The most commonly used ratio is OC:BC or OC/BC, but other
ratios include OC/EC, and OM/BC, where OM represents the total
mass of organic matter. Chapter 4 of this report provides OC/BC
and BC/PM2.5 ratios for a num ber of source categories.
770
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Several factors limit the value of the OC/BC ratio
as a short-hand for estimating the climate forcing
impact of a combustion source. Assuming that all
particle OC (including BrC) is scattering, regardless
of location, may underestimate the positive forcing
(warming) impact of a given source. Also, the OC/
BC ratio ignores other climate-relevant pollutants.
Sulfates, nitrates and secondary organic aerosols
(SOA) or their precursors, sulfur dioxide (SO2), NOX,
and VOCs, form additional light-scattering material
within the plume. Ramana et al. (2010) note that
the extent of BC-induced warming depends on the
concentration of both sulfate and OC. Further, the
aging process described in Chapter 2 induces optical
changes in an emitted particle mixture, including
coating of BC particles, leading to enhanced light
absorption (Lack and Cappa, 2010). These effects
are not captured by an OC/BC ratio. Finally, many
analyses that employ OC/BC thresholds for "net
warming effects" do not take into account other
effects, such as effects on precipitation and all the
indirect effects related to particle-induced changes
in clouds.
Despite the many limitations of OC/BC ratios,
they provide a simple way to evaluate potential
climate benefits and therefore continue to be used
to prioritize mitigation options. For the reasons
stated above, these ratios should serve only as a
approximate indicator of potential radiative effects
of categories of emissions sources. The specific
circumstances or policy goals should override
generic OC/BC rankings in cases where, for example,
emissions are affecting the Arctic, since even
mixtures that contain more reflective aerosols can
lead to warming over such light-colored surfaces.
In addition, OC/BC ratios are irrelevant to effects
that are shared among BC and other aerosols.
This includes precipitation or dimming effects,
and impacts on public health. For these types of
effects, mitigation strategies that reduce direct PM2.5
emissions or overall ambient PM2.5 concentrations
will provide the largest benefits, and the ratio of BC
to other constituents is far less important.
• Cost: Cost is a prime consideration in both
developed and developing countries. Regardless
of whether control technologies or alternative
strategies to reduce pollution are available,
policymakers are only likely to pursue such
approaches if the costs are limited and/or clearly
outweighed by the benefits of mitigation. For
some sectors and locations, BC control costs are
low in comparison to the public health benefits
that can be achieved (the most easily quantified
stream of benefits). However, control costs can
vary significantly depending on the technologies
being considered, and in many cases, particularly
in developing countries, there is a lack of
information about costs of BC control.
Existing Regulatory Programs: Countries
with existing regulatory programs and statutory
mandates may face legal constraints in terms
of the types of emissions reductions they
can pursue. There may also be technological
barriers and phase-in schedules that need to be
considered. For example, regulations that require
application of specific control technologies or
approaches may dictate the extent of emissions
reductions achieved. Many regulations are
phased in or become effective over an extended
time period. Controls on new mobile diesel
engines, for example, only reduce BC emissions
as new vehicles and engines are purchased and
deployed to replace older models. Policymakers
must work within existing regulatory constraints
in selecting mitigation initiatives.
Implementation Barriers: As policymakers
look across the range of available emissions
reductions, the other key factor to consider is
feasibility of implementation. This is perhaps
the most important consideration, because
there are some constraints that simply cannot
be overcome. Many examples exist for how
physical and political constraints hinder full
implementation of even those strategies for
which there is high confidence of large benefits.
Replacement of the 500-800 million traditional
solid-fuel burning cookstoves in developing
countries has been the subject of numerous
public and private campaigns, with one notable
large-scale success (China) and many cases of
limited success (Sinton et al., 2004). Only recently
has progress begun to accelerate, as illustrated
by the rapid growth in stove sales tracked by the
Partnership for Clean Indoor Air (see Chapter
10) and the emergence of the Global Alliance
for Clean Cookstoves. The historically slow
progress on this issue is attributable, in part, to
the scale of the problem and the difficulty of
implementing replacement programs in diverse
local environments.
Some studies have suggested altering the
timing of agricultural burning in the United
States to avoid transport to the Arctic during
the spring melt season (CATF, 2009a). However,
the seasonally dependent cycles of planting,
harvesting and pest-control (among other
considerations) make this a difficult strategy to
implement. Also, for stringent PM standards
for new vehicles, nonroad diesels, locomotives,
and commercial marine (other than ocean going
vessels), ultra low sulfur diesel fuel is required
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to enable use of diesel particulate filters. Many
countries already have regulations requiring such
fuels. As another example, there is near universal
agreement that retrofitting the existing diesel
fleet across the globe would reap tremendous
public health and climate benefits; however, this
kind of program would be very expensive. Cultural
practices also come into play. Residential heating
in the Arctic region has recently been identified
as a key contributor to BC impacts in the Arctic.
Nevertheless, both the rising cost of fossil fuel
in relation to wood and the cultural connection
to in-home wood heating makes a widespread
transition to pellet stoves (a recommendation
from the UNEP/WMO report) problematic.
• Uncertainty: In cases where there is a high
degree of uncertainty, policymakers may hesitate
to take mitigation actions. For example, if the
emissions mixture from particular sources is
not well characterized, or if the effectiveness
of control strategies is not well understood,
the benefits of pursuing reductions may be
questioned by decision-makers. This is particularly
true if the health and environmental co-benefits
of a strategy are limited, or if costs are high.
Under these circumstances, policymakers may
choose to postpone BC mitigation actions, or limit
those actions to a narrower set of controls with
more clearly defined costs and benefits.
All of these factors affect which mitigation options
are most desirable under different circumstances.
Clearly, the complexity of the decision-making
calculus requires the balancing of multiple
considerations simultaneously. In some cases,
these considerations may involve tradeoffs, and
policymakers will have to evaluate which options are
best given competing goals and considerations. The
following section illustrates how the different factors
in the decision-making framework can affect which
options are preferable under different circumstances.
7.6 Applying the Mitigation
Framework
To craft policy that addresses remaining BC
emissions, as well as anticipated increases in BC
emissions in some sectors and regions, policymakers
are faced with a complex set of considerations
and potentially competing goals. Optimizing
climate and public health benefits requires a broad,
multi-pollutant approach to BC mitigation that
includes looking at the entire suite of options and
evaluating them carefully to understand the full
range of possible benefits. Currently, most countries
implement a pollutant-by-pollutant approach to
air quality management, looking at each pollutant
in isolation and developing strategies targeting
each one. Though successful in many respects, this
approach may not result in the most effective or
efficient strategies for achieving multiple objectives.
As a result, many countries are moving toward
a multi-pollutant approach, where strategies
incorporate many pollutants by sector or by region.
However, most have yet to incorporate climate
pollutants and benefits into those strategies.
A recent case study of Detroit, Michigan conducted
by EPA explored how costs and benefits would
change if such a multi-pollutant framework was
designed (Wesson et ai., 2010). This case study
compared a traditional pollutant-by-pollutant
strategy (status quo) with one that sought to
maximize emissions reductions and risk reduction
for ozone, PM and selected air toxics (multi-
pollutant, risk-based). The study found that the
multi-pollutant, risk-based strategy produced over
twice the monetized benefits of the status quo.
Though the multi-pollutant, risk-based strategy
cost slightly more, it resulted in a much more
favorable benefit/cost ratio. This result was due, in
part, to a shift from reductions in secondary PM
to a strategy that achieved greater reductions in
primary PM, particularly in emissions from sources
affecting vulnerable and susceptible populations.
This is exactly the kind of shift in strategy that
could achieve BC benefits. This suggests that multi-
pollutant assessment can produce more cost-
effective strategies, and that targeting primary PM
emissions (with BC benefits) can have large health
benefits, especially in urban areas.
Policymakers facing choices among different BC
mitigation options can apply the 3-step framework
described above (defining goals, identifying
emissions reduction potential, and weighing
key factors) to evaluate the pros and cons of
different options and to identify the options that
best maximize co-benefits. In many cases, the
most desirable mitigation approaches will be
determined by the primary goal (or goals) combined
with constraints on resources, technologies, or
anticipated impacts. Several examples can help
illustrate how the factors in the framework affect the
attractiveness of different mitigation options.
In the United States, BC is managed as part of the
larger PM mitigation program designed to achieve
a suite of health and environmental goals. While
this generally means focusing on reducing total
PM mass, a greater emphasis on climate protection
as a goal in the PM program might require more
focus on reducing BC emissions within the overall
PM mixture. The largest domestic categories of BC
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Mitigation Overview: Designing Strategies for Public Health and Near-Term Climate Protection
emissions are mobile diesel engines and wildfires;
emissions from both sectors are large, so both
present some mitigation potential. However, there
are substantially different mitigation challenges
for each sector. Mobile diesel emissions are well
characterized and dominated by BC (as opposed to
other co-pollutants such as NOX, SOX, or OC); they
come from a defined set of controllable sources;
they can be controlled with known technologies with
clearly identified costs; they tend to be concentrated
in urban areas where they affect large populations;
and they can be regulated through centralized
federal programs. On the other hand, mobile
diesel emissions from U.S. sources are spread out
geographically and while they contribute to overall
atmospheric warming and have a large health impact
in urban areas, they are not as readily transported
to the Arctic. Controls on mobile diesel engines
can be expected to provide widespread health and
environmental benefits, and climate benefits in the
form of overall reductions in radiative forcing. These
global climate benefits will accrue in part to the
Arctic, helping to alleviate impacts (such as rapid
temperature increase) in that region. However, since
emissions within or near the Arctic appear to be
most closely linked to Arctic impacts (Quinn et al.,
2011), diesel reductions in the United States may
not be the most effective emissions reductions for
reducing impacts in the Arctic. They will have even
less impact on the Himalayas.
Wildfire emissions in the United States are less
well characterized in terms of volume and radiative
forcing potential, in part because they are dominated
by OC rather than BC. They are unpredictable and
extremely variable across time and space. They
are difficult to control with known technologies
or approaches, and often occur in locations far
removed from people. Implementation barriers are
much higher, and anticipated impacts of emissions
reduction strategies are much more uncertain. On
the other hand, wildfire emissions have been shown
to be one of the prime contributors to BC deposition
and atmospheric loading in the Arctic, and even
lighter-colored OC has been linked to warming in the
region (Quinn et al., 2011). Because the volume of
emissions from a single event is large and emissions
plume height enables long-range transport, wildfire
emissions can travel long distances and have impacts
far from the location of the fire. In the United States
in 2002, nearly 50% of wildfire emissions occurred
in Alaska, making them extremely important for
impacts on the Arctic.3 Therefore, if policymakers
3 Alaskan wildfire emissions exhibit significant interannual variability.
Emissions in 2002 were particularly high. Factors such as the timing,
extent, and location of the fire, the total volume (and type) of fuel
consumed, and the burning conditions all affect the fire emissions
and the net im pact on climate.
are seeking mitigation options that provide climate
benefits to the Arctic specifically, they might
be interested in attempting to control wildfire
emissions.
Outside of the United States, similar tradeoffs and
challenges exist. Agricultural burning in Africa is
one of the world's largest sources of BC. Despite
high OC co-emissions, this agricultural burning may
have substantial climate impacts because emissions
are lofted above deserts, which, like snow and
ice, are light-colored surfaces. However, because
much of the burning takes place in rural areas,
controlling these emissions would provide fewer
benefits for health than reducing emissions from the
transportation sector in larger African urban areas,
or reducing emissions from solid-fuel cookstoves.
Again, policymakers must consider a whole range of
factors in determining which options are preferable.
Some of the largest climate benefits of BC-focused
control strategies may come from reducing
emissions that affect the Hindu Kush-Himalayan-
Tibetan Plateau (HKHT), as well as other ice- and
snow-covered regions. The HKHT and Indian
subcontinent is also experiencing some of the most
dramatic impacts on precipitation from particle
pollution; these multiple climate effects mean
there is a tremendous opportunity in that region
to maximize climate benefits (including reduced
interference with natural precipitation patterns in
the region). An aggressive mitigation strategy could
reap substantial benefits directly to that region's
population, one of the world's largest. Here, many
of the variables in the framework align: a variety
of sources (cookstoves, diesel trucks, brick kilns)
are affecting both climate and health, and low-cost
technological solutions are available to help mitigate
these emissions. These technologies are discussed
further in the next several chapters, and some of
the key options for the region are highlighted in
Chapter 12.
7.7 Conclusions
Mitigating BC emissions depends on a clear
prioritization of goals and emissions reduction
opportunities. The challenge for policymakers
is clear: identifying feasible and cost-effective
mitigation strategies requires carefully weighing
a large number of factors, in many cases with
incomplete information. The challenge is rarely a
purely technological one. As the following chapters
will illustrate, many effective control technologies
are available to reduce BC emissions. These
control technologies can provide cost-effective BC
emissions reductions from key source categories,
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Chapter 7
and current studies (discussed in Chapter 6) suggest
that applying available technologies and strategies
will produce near-term climate benefits, especially at
the regional level. Available control technologies can
also provide substantial health and environmental
benefits. The challenge is identifying the strategies
that maximize benefits across all these categories.
Despite what is known and achievable, the path
forward is neither straightforward nor easy, and will
vary greatly by country and region of the world.
Nevertheless, as demonstrated in the previous
chapter, the potential benefits of action are large.
The next four chapters provide an overview of
BC mitigation options, including costs where that
information is available, in each of four major source
categories: mobile sources, stationary sources,
residential cooking and heating, and open biomass
burning. Policymakers can view these options
through the lenses discussed above and determine
which, if any, strategies are appropriate to reduce
BC emissions in their specific context. In Chapter
12, some of the clearest mitigation opportunities
based on current emissions, control technologies,
and expected benefits are discussed. These are
options that satisfy many of the key criteria that
decision-makers care about, as presented in the
mitigation framework above. They are options that
are most clearly linked to beneficial outcomes,
because of their high BC emissions reductions
potential, their importance for particular regions, or
the lack of constraints on implementation. However,
policymakers will need to adapt even these
important mitigation options to their particular local
and national circumstances.
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Chapter 8
Mitigation Approaches for Mobile
Sources
8.1 Summary of Key Messages
• In the United States, mobile sources accounted for
52% of total BC emissions in 2005, approximately
93% of which came from diesel vehicles or
engines. On a global basis, mobile sources
are responsible for approximately 19% of BC
emissions, with total mobile source emissions and
the percentage attributable to mobile sources
both significantly lower in developing countries.
• In the United States, new engine requirements
have resulted in a 32% reduction in BC emissions
from mobile sources between 1990 and 2005. As
vehicles and engines meeting new regulations are
phased into the fleet, a further 86% reduction in
BC emissions from mobile sources is projected
from 2005 to 2030, leading to a total decline of
90% in BC emissions between 1990 and 2030.
Such regulations have been effective in reducing
emissions of BC from on-road vehicles (mainly
diesel trucks), and nonroad diesel engines,
locomotives, and commercial marine vessels.
- Most of these reductions are concentrated
in the diesel fleet, and can be achieved via
application of diesel particulate filters (DPFs)
combined with ultra low sulfur diesel fuel.
DPFs typically eliminate more than 90% of
diesel PM and can reduce BC by as much as
99%.
- The cost of controlling PM2.5 from most types
of diesel engines is about $14,000/ton (2010$)
based on prior EPA rulemakings.
• Mobile source BC emissions in other developed
countries have been declining rapidly since the
1990s due to regulations on PM emissions from
new engines, mainly diesel trucks, and substantial
further emissions reductions are expected
by 2030 and beyond. Internationally, other
developed countries have and are continuing to
adopt emission standards (including those for
diesel engines with ultra low sulfur fuel) similar
to EPA emission standards, which also results in
harmonization of standards. However, standards
for new engines lag behind in some regions.
Of the on-highway and nonroad diesel engines
currently in operation in the United States, many
of which will remain in operation for the next 20
to 30 years, there are approximately 11 million
legacy fleet engines that are emitting PM at
elevated levels compared to new engines.
For policymakers seeking additional BC emissions
reductions beyond those that will be achieved as
a result of the new engine regulations already in
place, there are currently available, cost-effective
diesel retrofit strategies that can reduce harmful
emissions from in-use engines substantially.
- DPFs in a retrofit program for in-use vehicles
can reduce PM emissions by up to 99%,
at a cost of $8,000 to $15,000 for passive
DPFs, and $20,000 to $50,000 for active DPF
systems. However, not all engines are good
candidates for DPFs because of old age or
poor maintenance. Other cleaner engine
strategies include engine repowers, engine
upgrades, and replacement of the engine
(sometimes including the vehicle or piece
of equipment). EPA's National Clean Diesel
Campaign has provided grant funds to
support diesel engine retrofits, repowers, and
replacements.
- Other strategies to reduce emissions from
existing engines include improved fleet
maintenance practices, idle reduction
programs, advanced aerodynamics, more
fuel efficient tires and more efficient supply
chain management strategies, including
shifts in mode of transportation. EPA's
SmartWay Transport Partnership is designed
to encourage industry to adopt these
best practices for reducing emissions and
improving fuel economy.
- Internationally, retrofit programs present
significant financial and logistical challenges.
This is particularly true in developing
countries, where infrastructure is lacking to
assist with vehicle registration, inspection
and maintenance programs, technology
certification/verification programs, and
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Chapter 8
application of readily available technologies.
Vehicles in these regions tend to be older
and less well-maintained than in developed
countries, and the availability of low-sulfur
diesel fuel is limited. In addition, the costs of
DPFs may be prohibitive for some countries.
8.2 Introduction
A number of PM2.5 control strategies have proven
successful in reducing BC emissions from mobile
sources, which represent one of the most important
categories of BC1 emissions globally, especially
within developed countries (see Chapter 4). The two
principal strategies include: (1) emissions standards
for new vehicles and engines, with emissions
reductions occurring as the vehicle and engine
fleet turns over, and (2) controls or strategies that
reduce emissions from existing in-use engines, such
as diesel retrofits. In this chapter, these two major
strategies are explored, with emphasis on describing
the anticipated impact of these approaches on
emissions by 2030. It is important to note that these
strategies are complementary, and can be employed
simultaneously. The joint application of new engine
standards and controls on in-use engines has been
very successful in both the United States and Europe
in reducing direct PM emissions—including BC—
from mobile sources.2
Existing programs provide important insights
into achievable emissions reductions, costs, and
implementation challenges for new and existing
vehicles/engines in the mobile sector. Emphasis
is placed on programs and strategies which have
proven successful in the United States, including
both new vehicle/engine standards and programs
addressing in-use diesels such as EPA's National
Clean Diesel Campaign (NCDC), the SmartWay
Transport Partnership Program, and California's
mandatory diesel retrofit program. The chapter
discusses the impact of these approaches on
current and anticipated future emissions levels, and
1 As mentioned in Chapter 5, optical measurements of BC are limited
and vary depending on measurement technique. Measurements
of elemental carbon (EC) by thermal optical methods are more
widespread and consistent; mobile source emissions inventories
and information about control strategies for mobile sources usually
involve EC measurements. To ensure consistency in this report,
however, the term BC is used throughout.
2 Roughly 98% of the exhaust PM emitted from mobile sources is 2.5
microns or smaller in size. This is true for both diesel and gasoline
vehicles/engines. All exhaust particulate from mobile diesel sources
is commonly referred to as "diesel PM" and this convention is used
in this chapter. These emissions do not include secondary PM (SOA,
nitrates, sulfates) formed from mobile source emissions in the
atmosphere or tire and brake wear emissions.
describes the specific control technologies and
strategies involved, along with the cost of these
approaches. A close examination of such strategies
may offer insights into applicability of such
strategies elsewhere.
The main technology for reducing black carbon
emissions from diesel engines is the catalyzed
diesel particulate filter (DPF) discussed later in this
section. It is important to note that since DPFs are
made inoperable by fuels with high sulfur content,
mitigation of mobile source BC emissions depends
on the availability and widespread use of ultra low-
sulfur fuels (15 ppm sulfur). Typically, the low-sulfur
diesel fuel is in the marketplace about the same
time that the DPFs are introduced, although some
countries, particularly in the developing world, may
introduce low-sulfur fuel before adopting stringent
PM emission standards. The timing of ultra low-
sulfur fuel availability in different world regions is
discussed in this section, and in further detail in
Appendix 4.
8.3 Emissions Trajectories for Mobile
Sources
As discussed in Chapter 4, mobile sources remain
the dominant emitters of BC in developed countries.
In the United States, for example, mobile sources
were responsible for about 52% of BC emissions
in 2005, almost all of which (93%) came from
diesel vehicles or engines. If wildfire emissions are
excluded, then mobile sources account for 69% of
the 2005 domestic inventory. On a global basis,
mobile sources are responsible for approximately
19% of the BC (Bond et al., 2004) with total
emissions and percentage attributable to mobile
sources both significantly lower in developing
countries. A number of studies have projected that
these emissions are likely to increase globally in the
future, largely due to growth in the transportation
sector in developing countries (Streets et al.,
2004; Jacobson and Streets, 2009) (see Chapter 7).
However, mobile source BC emissions in developed
countries have been declining rapidly since the
1990s. Regulations on (PM) emissions from new
engines, particularly in the United States and
Europe, have been effective in reducing emissions
of BC from on-road vehicles (mainly diesel trucks),
and nonroad diesel engines, locomotives, and
commercial marine vessels, although Europe has
not currently adopted stringent locomotive and
commercial marine standards as the United States
has. Substantial emissions reductions are expected
over the next two decades and beyond.
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Mitigation Approaches for Mobile Sources
In the United States, new engine requirements
have resulted in a 32% reduction in BC emissions
from mobile sources between 1990 and 2005. As
vehicles and engines meeting new regulations are
phased into the fleet, a further 86% reduction in BC
emissions from mobile sources is projected from
2005 to 2030, leading to a total decline of 90% in
BC emissions between 1990 and 2030 as shown in
Table 8-1. Most of these reductions are concentrated
in the diesel fleet. For example, from 1990-2005,
there was a 30% decline in BC emissions from diesel
trucks. Due to new regulations, a further 95% decline
is projected in diesel truck BC emissions by 2030
(97% total decline since 1990). Other categories
of diesel engines, such as nonroad diesels (e.g.,
agricultural, construction equipment), commercial
marine diesels (excluding ocean going vessels),
and locomotives are also projected to have major
declines (75-92%) in BC emissions from 2005 to 2030
in the United States. BC emissions from gasoline
vehicles and nonroad gasoline engines, which are
much smaller sources of BC, are projected to decline
by 80% during 1990-2030 time period, with a 23%
reduction occurring from 2005-2030. Most of that
reduction will come from on-road gasoline vehicles
due to the use of catalysts that decrease PM.3'4
Considering only the emissions from U.S. mobile
sources occurring north of the 40th parallel in 2005,
EPA estimates there will be a substantial decline
of approximately 85% in these emissions by 2030
as well. As discussed in Chapter 4, emissions from
sources in northern latitudes are of particular
interest, due to the proximity of these emissions to
the Arctic and the greater likelihood of transport
to that sensitive region. However, the projected
decline in mobile source emissions north of the 40th
parallel does not reflect potential future increases
in emissions from marine freight transport that
may occur under future climate scenarios. The
total or seasonal loss of Arctic sea ice may result in
new marine trade routes through the Arctic. Such
3 Unlike the reductions for diesels, the reductions in BC from
gasoline engines occurred due to regulation of other pollutants
(such as hydrocarbons [HC], carbon monoxide [CO], and oxides
of nitrogen [NOJ) rather than regulation of PM itself. The use of
catalysts on these vehicles to decrease HC, CO, and NOX also results
in substantial PM and BC reductions. In general, BC emissions from
gasoline vehicles and engines have been less studied than those
from diesel engines.
4 Tire and brake wear are also considered to be mobile sources.
Emissions from these categories in the United States increased
from 1990 to 2030 due to increases in vehicle miles traveled (VMT).
Tire and brake wear are relatively minor sources of BC compared to
exhaust emissions (i.e., less than 1% of the total in 1990 but 4% in
2030) although they are larger from a PM standpoint. Importantly,
BC accounts for 22% of PM em issions from tire wear. At present,
there are no EPA emission standards for either tire or brake wear PM
emissions.
developments could potentially result in greater
emissions in the Arctic, with greater potential
for deposition on remaining ice. U.S. emissions
inventories currently contain no projections of
these potential future emissions in the Arctic area.
However, some studies have been done of emissions
from shipping and aircraft in the Arctic area (Corbett
et al., 2010; Wilkerson et al., 2010).
Table 8-1 shows the emissions reductions in BC (as
well as PM2.5 and OC) going from 1990 through
2030 for various mobile source sectors which are
discussed in the following sections. The basis for
the emissions inventories here is discussed in the
mobile source section of Appendix 2. The numbers
are based largely on the MOVES and NONROAD
models, which represent EPA's projections for
emissions reductions that will occur as a result
of the engine and tailpipe emissions regulations
already promulgated by EPA, but do not include any
additional emissions reductions that would occur
as a result of engine retrofits or replacements. Also,
Figure 8-1 shows the reductions in BC graphically
from 1990 through 2030.
8.4 New Engine Standards in the
United States
In the United States, PM emissions standards for
new mobile source engines are being phased in
across different sectors between 2007 and 2020,
mostly for diesel engines. These standards will
lead to the large reductions in mobile source
emissions of BC illustrated in Table 8-1.5 The
realized reductions depend on the rate of fleet
turnover—i.e., the rate at which older vehicles and
engines are replaced with new vehicles that comply
with the latest emissions standards. The rate of fleet
turnover depends heavily on the type of vehicle or
engine, with on-road engines such as passenger
cars and light-duty trucks being replaced more
frequently than some other types of mobile sources,
such as nonroad equipment. The state of California
has its own diesel PM standards as promulgated by
the California Air Resources Board (CARB). These
standards are, in general, similar if not identical
to the Federal standards. CARB also has its own
gasoline PM standards. A detailed list of the mobile
source PM standards is contained in Appendix 5.
The emission standards and/or control technology
cited below to reduce PM (and thus BC) emissions
do not include programs such as increased use of
5 EPA models the cumulative reductions for each category of mobile
sources attributable to all past and current standards promulgated
for that category rather than modeling the reduction for a
particular standard.
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Chapter 8
Table 8-1. Mobile Source BC, OC, and PM2.5 Emissions 1990-2030 (short tons). (Source: U.S. EPA)
Source Category Year % Change
BLACK (ELEMENTAL) CARBON
Onroad gasoline
Onroad diesel
Tire
Brakewear
Nonroad gasoline
Nonroad diesel
Commercial Marine (C1 & C2)
Commercial Marine (C3)
Locomotive
Aircraft"
Total BC Emissions (Mobile)
1990
69,629
219,958
809
290
5,420
148,537
22,122
1,262
19,317
283
487,628
2005
14,510
153,477
1,198
475
5,444
112,058
21,652
1,681
22,495
410
333,400
2020
9,538
28,175
1,435
569
4,702
31,254
11,595
864
11,349
457
99,940
2030
10,027
7,615
1,720
682
5,174
9,356
5,440
1,306
5,684
553
47,557
1990-»2005
-79%
-30%
48%
64%
0%
-25%
-2%
33%
16%
45%
-32%
2005-»2030
-31%
-95%
44%
44%
-5%
-92%
-75%
-22%
-75%
35%
-86%
ORGANIC CARBON
Onroad gasoline
Onroad diesel
Tire
Brakewear
Nonroad gasoline
Nonroad diesel
Commercial Marine (C1 & C2)
Commercial Marine (C3)
Locomotive
Aircraft"
Total OC Emissions (Mobile)
262,065
66,056
1,734
1,191
37,613
33,872
5,045
4,734
4,405
1,372
418,088
59,657
44,423
3,060
2,321
46,734
30,618
4,937
6,303
5,130
1,988
205,172
43,711
14,883
3,678
2,790
41,137
9,759
2,772
8,644
2,659
2,217
132,252
47,421
10,580
4,407
3,343
45,424
3,891
1,710
13,060
1,507
2,682
134,025
-77%
-33%
76%
95%
24%
-10%
-2%
33%
16%
45%
-51%
-21%
-76%
44%
44%
-3%
-87%
-65%
107%
-71%
35%
-35%
DIRECT PM2 5
Onroad gasoline
Onroad diesel
Tire
Brakewear
Nonroad gasoline
Nonroad diesel
Commercial Marine (C1 & C2)
Commercial Marine (C3)
Locomotive
Aircraft"
Total PM2 5 Emissions (Mobile)
335,205
290,478
3,678
11,129
54,198
192,905
28,730
42,082
25,087
2,178
985,671
75,924
208,473
5,325
17,801
55,834
145,289
28,119
56,028
30,910
3,156
626,859
54,682
43,698
6,450
21,559
49,000
46,310
15,789
14,407
15,145
3,519
270,559
59,106
18,765
7,727
25,830
54,078
18,463
9,741
21,767
8,584
4,257
228,318
-77%
-28%
45%
60%
3%
-25%
-2%
33%
23%
45%
-36%
-22%
-91%
45%
45%
-3%
-87%
-65%
-61%
-72%
35%
-64%
Non landing and take-off (LTO) emissions not included;
require funding for implementation are not included in
also, planned technology and operations improvements that
the forecast.
electrification (either for light-duty vehicles using
hybrids or electric vehicles or, more importantly,
truck stop electrification which reduces idling of the
diesel truck engine and use of auxiliary power units
on heavy-duty trucks which are typically small diesel
engines). They do not include benefits from reduced
idle programs or other transportation control
measures (such as reduced commuting, increased
778
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Mitigation Approaches for Mobile Sources
600,000
(a)
500,000
400,000
300,000
200,000
100,000
450,000
Aircraft
Locomotive
Commercial Marine (C3)
Commercial Marine (Cl &
C2)
• Nonroad diesel
• Nonroad gasoline
• Brakewear
• Tire
• Onroad diesel
• Onroad gasoline
Aircraft
Locomotive
D Commercial Marine (C3)
Commercial Marine (Cl & C2)
Nonroad diesel
• Nonroad gasoline
• Brakewear
• Tire
• Onroad diesel
• Onroad gasoline
1990 2005 2020 2030
1990 2005 2020 2030
1,200,000
(C)
1,000,000
800,000
g 600,000
I
a
400,000
200,000
Aircraft
Locomotive
Commercial Marine (C3)
Commercial Marine (Cl &
C2)
• Nonroad diesel
• Nonroad gasoline
• Brakewear
• Tire
• Onroad diesel
• Onroad gasoline
Figure 8-1. Estimated Changes in Emissions of (a) BC,
(b) OC, and (c) Direct PM2 5 from Mobile Sources in the
United States, 1990-2030. Estimates of the number of tons
of emissions reduced from each mobile source category
are reported in Table 8-1. (Source: U.S. EPA)
1990 2005 2020 2030
use of mass transportation, increased bicycling/
walking). These types of programs are discussed
more generally in a later section of this chapter.
The reductions in BC also do not consider how BC
would be affected by future fuel economy standards
such as those for light-duty vehicles (which are
mostly gasoline-powered and thus a smaller source
of BC emissions) and diesel vehicles (which are
mostly heavy-duty trucks and a larger source of BC
emissions). EPA has issued light-duty vehicle fuel
economy standards effective for the 2012-2016
model years. EPA also just issued final rulemaking
for heavy-duty vehicle fuel economy standards for
the 2014-2018 model years. Additional fuel economy
improvements for light-duty vehicles for model years
2017-2025 have recently been proposed.6 Basically,
these standards will not increase BC emissions.
These rulemakings and other forces in general
will result in changes in vehicle technology. The
introduction of and increased use of electric
vehicles is certainly already occurring. There have
been several studies (Jacobson and Delucchi,
2011; Delucchi and Jacobson, 2011) examining
alternative energy sources including one on
providing worldwide energy (for electric power,
transportation, heat/cooling) by wind, water, and
sunlight on a widespread basis in the 2030-2050
time frame. These alternative power sources could
' See http://www.epa.gov/otaq/climate/regulations.htm.
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Chapter 8
greatly reduce emissions of PM and BC. Also, it will
be important to determine the effect from increased
use of biofuels on BC emissions, which is currently an
area of significant uncertainty.
8.4.1 On-road and Nonroad Diesel Engines
Diesel PM, as it exits the engine, is 70-80% BC for
the pre-2007 model year diesel trucks and current
diesel nonroad engines (excluding commercial
marine oceangoing vessels which are discussed
separately). The main source of diesel PM has
traditionally been heavy-duty diesel trucks with
gross vehicle weights from 8,501 to 80,000 Ibs. The
first standards controlling diesel PM for on-road
engines were standards for visible smoke (which has
some correlation with PM) effective with the 1970
model year followed by increasingly stringent PM
mass standards starting with the 1988 model year.
For the 2007 vehicle (engine) model year, stringent
emission standards of 0.01 g/BHP-hr (grams
per brakehorsepower/hour - a standard unit for
emissions from heavy-duty mobile source engines)
became effective for heavy-duty diesel engines,
which represents over 99% control from a pre-
control diesel engine in the 1970 time frame.7
As a result of these standards, BC emissions have
been dramatically or even preferentially reduced as
the major PM constituent.8 To meet these stringent
PM standards, virtually all new on-highway diesel
trucks in the United States, beginning with the 2007
model year, have been equipped with DPFs. DPFs
typically eliminate more than 90% of diesel PM
and can reduce BC by as much as 99%. The type of
DPFs typically used on new model year vehicles are
called "wall flow" filters with a catalyst coated on a
ceramic monolith with the exhaust flowing through
the filter walls trapping the PM and allowing the
exhaust gases to flow through. The trapped PM is
then oxidized by reaction with compounds such as
oxygen and nitrogen dioxide on the catalyst surface.
This technology preferentially removes solid particles
7 EPA's emissions standards for heavy-duty diesel trucks have
always been engine standards since the same engine can be used
in a wide variety of truck chassis bodies with many of these bodies
manufactured by companies different from those who manufacture
the engines. For light-duty vehicles and trucks (trucks up to 8,500 Ibs
gross vehicle weight), the emission standards in g/mile apply to the
car/truck itself.
8 Ultrafine particles (generally those smaller than about 0.10 microns
in size) from pre-2007 diesel engines generally comprise primarily
BC, OC, metals, and sulfates. DPFs preferentially reduce BC, OC,
and metals. Also, the use of ultra low sulfur diesel fuel reduces total
sulfate emissions (and emissions of ultrafine sulfate PM). Recent
work shows that DPFs reduce particle number (an indicator of
ultrafines or nanoparticles) by up to 90-99% based on em issions
characterization with four 2007 heavy duty diesel engines. See
Khaleketal. (2009).
such as BC. BC emissions from the heavy-duty
diesel truck fleet have been reduced by 30% from
1990-2005, and EPA projects that the application
of DPFs will result in a further 95% reduction by
2030, from 153,477 tons to 7,615 tons. EPA's earlier
rulemakings concluded that use of DPFs separate
from the overall emission control system could
result in a minimal fuel economy penalty (-1%) due
to additional pumping work to force the exhaust
gases through the DPF at high engine loads, but that
the overall fuel economy impact would be neutral
due to optimization of the complete emission
control system. This was one of the primary reasons
the Agency took such a systems approach. Now
that the heavy-duty on-highway program is fully
phased-in, some manufacturers are claiming a 5-6
percent fuel economy improvement through the use
of integrated emission control systems. Additionally
EPA and NHTSA projected that these overall
optimized emission control systems could be further
improved as part of the technology packages engine
manufacturers are projected to use to comply with
the Agencies' recently finalized Heavy-Duty Fuel
Efficiency and Greenhouse gas rulemaking.
Corresponding national PM emissions standards
of 0.01 g/mile took effect for U.S. passenger cars
(and light-duty trucks) from 2004-2006. These
"Tier 2" standards apply to both gasoline and diesel
light-duty vehicles, although there are very few
diesel passenger cars in the United States (unlike in
Europe where diesel passenger vehicles are used
extensively).
Nonroad diesel engines also emit a significant
amount of BC. EPA's first emission standards
for PM for these engines began in 1996. Recent
rules issued in 2004, to be effective with the 2012
calendar year for newly manufactured engines,
will result in widespread use of DPFs with dramatic
reductions (~ 99% from a pre-control engine) in
PM and BC. These standards will be fully phased
in around 2015 for new model year nonroad diesel
engines but will be phased into the fleet some years
later with fleet turnover. EPA's latest version of the
NONROAD model calculates the effect of all of these
regulations, including those resulting in use of DPFs.
EPA calculates a 92% decrease in emissions between
2005 and 2030, from 112,058 tons of BC in 2005
to 9,356 tons in 2030, despite substantial expected
growth in use of these engines over this time period.
Cumulatively, this will be a 94% decrease from 1990
to 2030.
A general note is that the recent down turn in the
economy (not accounted for in these projections)
can result in lower fleet turn-over than seen
historically for on-road light-duty vehicles and
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trucks. This can also be an issue with nonroad
engines. These changes by themselves would
increase emissions since increased numbers of older
vehicles or engine are being used. Also, a shift in
travel patterns and freight movement can occur,
such as altered use of intermodal freight facilities.
Economic downturns may also reduce total usage
for both on-road and nonroad vehicles, which
would reduce total emissions. Similarly, increases
in fuel prices and land-use patterns will affect
transportation patterns. Also, any shift in travel
patterns and freight movement such as altered
use of intermodal freight facilities would affect BC
emissions. Finally, it is important to note that the
total emissions reductions achieved will depend on
the extent to which older vehicles/engines officially
retired from service are still utilized for limited
purposes in the United States or are exported to
other countries (especially in Central and South
America) for continued use.
As mentioned briefly in the introduction to
this chapter, an important prerequisite for the
application of DPFs is a switch to low-sulfur fuel.
Low-sulfur fuel is needed, and has been required in
the United States by regulation, to preserve catalytic
activity of the emission control system, which is
poisoned by sulfur. In issuing diesel PM regulations
for on-road heavy-duty vehicles, nonroad diesels,
and commercial marine (categories 1 and 2)/
locomotives, EPA determined that the emission
standards being required could be met only with
use of ultra low sulfur diesel fuel. Specifically, sulfur
interferes with the ability of the DPF to passively
regenerate. For NOX control with urea selective
catalytic reduction (SCR), sulfur compromises low-
exhaust temperature NOX reduction performance.
Fuel sulfur also results in sulfate PM due to catalytic
oxidation of sulfur oxides over the DPF, which
increases PM. Noncatalytic diesel particulate filters
that would be compatible with higher sulfur diesel
fuels are harder to regenerate (i.e., removal of
accumulated diesel particulate in the filter) and are
not as effective for PM control. They also do not
control the organic fraction of PM as effectively
and, thus, do not meet stringent PM standards (U.S.
EPA, 2001). Such filters though could be among
possible control technologies for larger commercial
marine diesels (category 3) which use heavy fuel oil
instead of conventional diesel fuel; these engines are
discussed later.
EPA first regulated sulfur content in on-road diesel
fuel to 500 ppm in 1993, resulting in typical fuel
sulfur levels of about 300 ppm. Prior to that, the
sulfur level in on-road diesel fuel was about 2,000
ppm. In 2006, the sulfur level was limited to 15
ppm for on-road diesel fuel and has been reduced
gradually in nonroad diesel fuel, first to 500 ppm in
2007 for all categories except ocean-going vessels,
and, starting in 2010, to 15 ppm for most categories.
In the case of locomotive and marine diesel fuel
(for categories 1 and 2 marine diesel), this second
step will occur in 2012. Thus, all highway diesel
vehicles and nonroad engines in the United States
must now or will soon operate on "ultra-low sulfur
diesel" (ULSD). Typical in-use fuel sulfur levels are
about 7 ppm. Of course, as discussed later, fuel for
the larger C3 marine (such as heavy fuel oil, HFO)
diesel (ocean-going) engines has significantly higher
sulfur levels and would not be suitable for diesel
particulate filters.
It is important to note that the net climate impact
of the application of DPFs will be offset somewhat
by the necessary co-emissions reductions in sulfate,
which is reflecting (cooling).9 Also, while diesel PM
from pre-2007 engines has a high level of BC in
PM, it also has some OC (about 22%), which is also
greatly reduced by the DPF in later model years. The
net climate impact of the application of DPFs will
be affected by these reductions in OC emissions.
Still, given the predominance of BC in diesel exhaust
(70-80%), emissions reductions from this source
category have a strong likelihood of providing
climate benefits.
The EPA nonroad diesel rule10 issued in 2004
provides an aggregate cost estimate for controlling
PM emissions using DPFs on new engines of about
$14,000 per ton ($2010). This cost figure includes
the additional cost of ULSD fuel, engine costs, any
changes in maintenance costs, and equipment costs.
As shown in Table 8-2, similar cost estimates were
developed in 2001 for the Heavy-Duty Diesel Rule
9 The 15 ppm sulfur limit greatly reduces SOX emissions, some of
which convert to sulfate in the ambient air. For exhaust emissions
of sulfates, the situation is more complicated since a typical
conversion rate of SO2 to sulfate for diesel engines without DPFs
is about 2% but increases to about 50% for vehicles/engines with
DPFs. Due to the dramatic reduction in diesel fuel sulfur, there is
still some reduction in sulfate emissions from vehicles/engines with
DPFs and 7 ppm diesel fuel sulfur versus vehicles/engines without
DPFs using fuel meeting the 500 ppm limit, which results in a typical
sulfur level of 200-300 ppm. A 50% conversion of SOX to sulfate
with the typical 7 ppm fuel sulfur level results in less exhaust sulfate
(about 35% less) than from an older pre-trap diesel using fuel with
the 200-300 ppm sulfur levels.
10 Control of Emissions of Air Pollution from Nonroad Diesel Engines
and Fuel. Federal Register: June 29, 2004 (Volume 69, Number 124).
See specifically, Final Regulatory Analysis: Control of Emissions from
Nonroad Diesel Engines, EPA420-R-04-007, Chapter 8, Table 8.7.1,
page 33, May 2004 (http://www.epa.gov/nonroaddiesel/2004fr.
htm#ria).
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Table 8-2. Cost Estimates for Particulate Matter Controls on New Diesel Engines (2010$), based on
Recent U.S. EPA rulemakings. These costs include the additional cost of requiring Ultra Low Sulfur Diesel
Fuel. (Source: U.S. EPA)
Estimated Cost (2010$) Per Ton PM2 5 Reduced
Rule 1 :
N PV, 3% rate N PV, 7% rate
Heavy-Duty Diesel Rule (2001)
Nonroad Diesel Rule (2004)
Locomotive/Marine Rule (2008)
$16,652
$13,762
$8,579
$19,216
$14,461
$9,778
for on-road11 and the 2008 rule controlling emissions
from locomotive and marine diesel engines.12 It
is important to note that the values reported in
Table 8-2 are adjusted from the original values
developed by EPA to 2010$ as a function of GDP to
ensure consistency with other costs presented in
this Report. A large fraction of the cost is due to the
requirements for ultra low sulfur diesel fuel.
It is important to note that the controls applied
under these regulations affect multiple pollutants,
not just BC. At this time, there is no methodology
to allocate these costs specifically to BC or other
PM components but it is useful to note that for
these diesels the BC is the largest PM component.
Furthermore, the analyses conducted during
the 2001-2008 time frame utilized the best cost
information available at that time, as well as
emissions reductions (total tons reduced) based
on EPA's then-current emissions models. Since
then, the emissions models have changed so that
the reductions estimated in the earlier rulemaking
analyses would be somewhat different today. The
magnitude of the reductions was determined doing
emissions inventory estimates for given years both
with and without the standard being considered in
effect. One cannot obtain the tons reduced by given
standard just from emissions inventory data for a
given year compared to another year since the total
11 Control of Air Pollution from New Motor Vehicles: Heavy-Duty
Engine and Vehicle Standards and Highway Diesel Fuel Sulfur Control
Requirements, Final Rule. Federal Register: January 18, 2001 (Volume
66, Number 12). This rule applies to 2007 and later model-year
heavy duty diesel on-road engines. See specifically, Regulatory
Impact Analysis: Heavy-Duty Engine and Vehicle Standards and
Highway Diesel Fuel Sulfur Control Requirements; Chapter VI, Table VI
F-4, page VI-19, January 2001 (http://www.epa.gov/otaq/highway-
diesel/regs/ria-vi.pdf).
12 Control of Emissions of Air Pollution from Locomotive Engines
and Marine Compression-Ignition Engines Less than 30 Liters per
Cylinder; Republication. Federal Register: June 30, 2008 (Volume
73, Number 126). See specifically, Regulatory Impact Analysis:
Control of Emissions of Air Pollution from Locomotive Engines and
Marine Compression Ignition Engines Less than 30 Liters Per Cylinder,
Chapter 5, Table 5-67, page 5-98, June 2008 (http://www.epa.gov/
oms/regs/nonroad/420r08001a.pdf).
reduction reflected in the inventory from one year
to another is the result of all the standards in place
(and vehicle/engine turn over) for all mobile sources
rather than just a single standard for a particular
category. Also, the inventory and cost numbers used
in these calculations have not been updated since
they were obtained. In the absence of new analysis,
the $14,000 cost/ton (the average costs in Table
8-2) is the best available EPA information for control
of diesel PM from newly manufactured on-road
vehicles and nonroad engines meeting EPA emission
standards. The total costs and benefits of these
regulations are discussed separately in Chapter 6. As
an aside, EPA cost estimates made in rulemakings
tend to be higher than the actual cost due to
improvements in technology to meet the standard
that were not considered when the rule was issued
(Anderson and Sherwood, 2002).
8.4.2 On-road and Nonroad Gasoline
Engines
On-road gasoline PM emissions have decreased
dramatically over the years, especially with the use
of catalysts and unleaded gasoline starting with
the 1975 model year vehicles. For example, PM
emissions for a typical car using leaded gasoline in
1970 were about 0.3 g/mile compared to emissions
from current vehicles with unleaded fuel of about
0.001 g/mile, a reduction of over 99% (Coordinating
Research Council, 2008). While BC emissions were
not usually measured in the PM from cars in the
1970s, some limited measurements suggest that
BC made up about 10-20% of the PM at that time,
compared to about 20% of PM mass in 2005. Thus,
the per-vehicle PM reductions since 1970 have
resulted in a substantial reduction in BC emissions.
Most of this BC comes under "rich" operating
conditions (where there is insufficient air for full
combustion, such as during cold-start or high load
conditions). EPA's most recent modeling indicates
that BC emissions from on-road gasoline engines
have declined 79% since 1990, from 69,629 in 1990
tons to 14,510 tons BC in 2005, and will decline a
further 31% by 2030 (to 10,027 tons).
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Under the Tier 2 exhaust regulations mentioned
above for light duty vehicles (passenger cars and
light-duty trucks), EPA set a PM emissions standard
for both gasoline and diesel vehicles at 0.01 g/mile
starting in 2004, with full phase-in for all light-duty
vehicles (including light-duty trucks) in model year
2009. When the Tier 2 rules were promulgated,
EPA estimated that a total of 36,000 tons of PM2.5
would be reduced in the year 2030 from these
standards (versus not having these standards)
using the emissions models available at that time
(U.S. Environmental Protection Agency, 1999). Prior
exhaust standards from the 1990s and earlier also
have helped reduce PM. While these regulations do
not limit PM directly, they resulted in better control
of air/fuel ratio and improved catalyst formulations
to meet HC, CO, and NOX emissions standards, all
of which affected PM emissions levels. Because
the regulations were targeted at other pollutants,
however, EPA has not calculated a cost for the
resulting PM reductions specifically.
It should be noted that most current technology
vehicles now emit below the Tier 2 PM standard
by a significant margin. However, a relatively new
technology, gasoline direct injection (GDI), is being
utilized for a number of reasons such as improved
fuel economy and performance. GDI engines differ
from conventional fuel injected engines in that the
fuel is injected directly into the cylinder (like in a
diesel engine) rather than at the intake port. GDI
vehicles are expected to constitute a major part of
the new vehicle fleet in the coming years and may
be 90% of new vehicle sales in model year 2016.
The specific technology for injecting and guiding
the gasoline spray into the engine coupled with the
catalyst may have an impact on the magnitude of
the PM emissions. Recent studies performed by EPA
determined that some "wall guided" GDI engines
perform slightly worse than currently produced "port
fuel injection" (PFI) engines with respect to PM but
that "spray guided" GDI engines perform on par with
PFI engines. Indications are that most manufacturers
utilizing GDI technology will be migrating to "spray
guided" designs, but regardless EPA anticipates
future GDI designs will perform on par with or better
than current technology.
CARB has issued a preliminary discussion paper
discussing the option of tightening the PM mass
standard effective for the 2015 model year (California
Air Resources Board, 2010). The present CARB
PM standard (LEV II) is 0.01 g/mile, which is also
the EPA emission standard. The possible standard
presented in the discussion paper is 0.003 g/mile
starting in 2017. A 0.001 g/mile standard is being
considered starting for the 2025 model year. CARB
had also considered a standard specifically for BC,
but announced at its November 2010 LEV III (Low
Emission Vehicle) workshop that it would not set
such a standard.
Nonroad gasoline engines are either 2-stroke
engines (where lubricating oil is mixed into and
burned with the gasoline) or4-stroke engines. The
2-stroke engines are smaller engines and tend
to be used more in lawn and garden equipment,
such as handheld string trimmers; they have also
been used in lawn mowers and snow blowers. They
can also be used in recreational marine, although
most engines there are now 4-stroke engines.
The 4 stroke engine is used in equipment such
as lawn mowers, small generator sets, industrial
equipment, and recreational equipment such as
marine engines. These engines emit significant PM
mass, especially the 2-stroke engines, where the
PM has a large contribution from the lubricating oil.
They can also be used in larger equipment such as
farm and construction equipment although, here,
the dominant engine type is diesel. EPA estimates
that BC emissions from nonroad gasoline engines
will decline approximately 5% (from 5,444 tons to
5,174 tons) between 2005 and 2030, largely due
to changes needed to meet standards for volatile
organic compounds (VOC), CO, and NOX emissions
standards being applied to several categories of
nonroad gasoline engines, which will also reduce
PM. Current information, which needs to be
updated, used in EPA air quality modeling suggests
that BC is approximately 10% of PM mass with the
same number being used for both 2-stroke and
4-stroke engines even though 2-stroke engines have
oil added to gasoline. PM emissions from nonroad
gasoline engines, particularly the 2-stroke engines,
have been characterized far less thoroughly than
emissions from on-road gasoline vehicles, and EPA's
estimates of BC emissions are highly uncertain.
EPA places a high priority on obtaining better BC
emissions data from both 2-stroke and 4-stroke
nonroad gasoline engines.
8.4.3 Other Mobile Sources - Commercial
Marine Vessels, Locomotives, and Aircraft
Locomotives have used diesel (diesel electric)
engines (both 2-stroke and 4-stroke engines)
predominantly since the 1950s. EPA has
implemented several tiers of emission standards
for PM for these engines with the most recent set
of standards to be effective in 2015. These newest
standards will result in the use of DPFs on new
locomotives which preferentially reduce BC. In
addition, national emission standards require that
older locomotives that are remanufactured must be
certified to more stringent emission standards than
their prior certification level.
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Commercial marine vessels are classified as Cl, C2,
and C3 based on engine size. Cl marine engines
are similar in size (less than 5 I/cylinder or for some
categories less than 7 I/cylinder) to those used in
construction/farm equipment. C2 marine engines
(between 5 or 7 and 30 I/cylinder) are similar to
locomotive diesels. The C3 engines (greater than
30 I/cylinder vessels) are similar to those used in
some power plants and are used in ocean-going
vessels. The most recent set of emission standards
for these engines will result in most new Cl and
C2 commercial marine engines having DPFs
starting in 2014. Ultra low sulfur diesel fuel is being
required for these engines. For these engines,
there will be a dramatic drop in PM emissions and
an even more dramatic drop in BC emissions. Like
locomotives, older marine diesel engines must
be certified to more stringent emission standards
upon remanufacturing, compared to their previous
certification level. The level of the standards to which
these remanufactured engines must be certified
varies depending on engine type and year of
manufacture for the original engine.
PM emissions from C3 engines comprise mainly
sulfate (about 75%) and relatively little BC (can be
less than 1% although as discussed in Appendix 2
this percentage can vary significantly). Due to recent
work with the International Maritime Organization
(IMO), there will be large reductions in the higher
sulfur level of the fuel (largely bunker diesel fuel
composed of especially high molecular weight,
even solid, hydrocarbon compounds) used in these
engines (see Appendix 4). As this sulfur level is
reduced, PM will be greatly reduced but BC levels
are expected to stay roughly the same on a per-
vessel basis and will constitute a larger percentage
of the PM emissions. There is some increase in BC
emissions from 2005 due to an increase in usage of
these vessels. Though C3 marine is responsible for
less than 1,000 tons of BC emissions for the entire
United States, there is some concern that emissions
from these vessels could have disproportionate
impact on the Arctic, especially if Arctic marine
traffic increases as shipping lanes open due to ice
melt in the region. Additional BC emissions data and
modeling/deposition studies are needed to clarify
the impact of C3 marine vessels.
C3 marine usually uses heavy fuel oil (HFO), which
can be solid at room temperature (and is heated
before going into the engine), rather than the
conventional distillate diesel fuel used by C1/C2
commercial marine and other nonroad diesels. HFO
contains higher molecular weight hydrocarbon
compounds than conventional distillate diesel fuel.
This affects the characteristics of the PM emissions.
As is also discussed in Appendix 4, HFO contains
higher fuel sulfur levels and cannot be used with
DPFs.
There has been limited research into the BC
emissions from aircraft. Additional characterization
of aircraft emissions would help to improve
understanding of BC emissions from aircraft,
although there is sufficient information to develop a
PM inventory and an initial BC and OC inventory.
In general, therefore, additional emissions
information for commercial vessels, locomotives
and aircraft would improve characterization of BC,
since present data are limited, and it is difficult to
determine how much BC will be reduced by the PM
standards affecting these sources.
8.5 New Engine Standards
Internationally
Heavy-duty on-road diesel vehicles represent
the predominant mobile source of BC in most
areas of the world, although nonroad diesel (and
locomotives and commercial marine) can also be
significant. Given the importance of diesel engines
internationally, use of DPFs to reduce PM2.5 will also
result in large reductions in BC from the global
mobile source sector. Some countries have already
made significant progress in this area and have
introduced diesel PM standards (mainly for on-
road vehicles) which effectively reduce BC. While
broad-scale application of DPFs is an attractive
option to reduce global emissions, this is dependent
on simultaneous use of ULSD fuel. Many other
developed countries in Europe and Asia have already
adopted low-sulfur fuel requirements. As a result, BC
emissions from mobile sources are declining in many
regions, especially in Europe and Japan. However,
many developing countries have not yet switched to
low-sulfur fuel, and PM emissions controls are less
common. Each of these issues is discussed further,
below. In general, the U.S. experience controlling
diesel PM and, thus, BC provides a good template
for international control programs.
8.5.1 International Regulations of Diesel
Fuel Sulfur Levels
As noted above, the availability and widespread
use of low-sulfur fuels is a critical prerequisite to
effective BC control from mobile sources. Like the
United States, Canada, Japan, and the European
Union have adopted strict controls on on-road
diesel fuel sulfur levels. Many countries have also
adopted regulatory standards for reducing sulfur
levels in on-road diesel fuel to levels needed to
enable low-emission vehicle technologies. In other
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regions, however, reductions in the sulfur content
of fuel lag behind. This effectively constrains BC
emissions reductions in these countries, since higher
sulfur fuels would prevent DPFs from functioning
properly, even if they were applied.
The United Nations Environment Programme's
(UNEP) Partnership for Clean Fuels and Vehicles
(PCFV), founded at the World Summit on Sustainable
Development in 2002, promotes low sulfur fuels
and cleaner vehicle standards and technologies.
This partnership has over 100 members from the oil
and gas industry, engine and retrofit manufacturers,
government agencies, and environmental NGOs.
Currently, the PCFV is conducting a low sulfur
campaign with a call for global adoption of 50 ppm
sulfur gasoline and diesel. The implementation of 50
ppm sulfur programs would allow countries to begin
to deploy DPFs, which would produce significant
reductions in PM2.5 and BC. However, the U.S. EPA
believes a further reduction to sulfur levels at or
below 15 ppm is needed for DPFs to function for
their intended lifetime. Further detail on the diesel
sulfur reduction activities of countries outside the
United States, Canada, Japan, and the European
Union is provided in Appendix 4. Most of the actions
underway in other countries focus on fuels for on-
road vehicles. Sulfur limits for nonroad diesel fuel
are also needed on an international basis to facilitate
BC control. It is important to note that the cost to
provide the ULSD fuel will vary from one country
to another depending on fuel supplies and refinery
capabilities. Thus, while the benefits of low sulfur
fuels and advanced emission control technologies
far outweigh the costs, the often substantial upfront
costs of upgrading existing refineries presents a
challenge for many governments.
The global community has also been working
to reduce the sulfur content of fuels used in
marine vessels. Currently, the IMO has established
requirements for the sulfur content of bunker type
fuel used in C3 marine vessels on both a global
basis and for an Emission Control Area (ECA) in
specific target years (U.S. EPA, 2010g). However,
these requirements are designed to reduce sulfate
emissions, rather than to enable use of DPFs, and
even the cleanest fuel on this schedule (1,000 ppm
sulfur within the ECA by 2015) would not enable use
of DPFs (see Appendix 4).
8.5.2 Standards for New Engines Outside
the United States
Many other countries have adopted PM emission
standards for new engines. Most of these standards
affect on-road engines, and the rigor of these
standards and the time for phase-in of new engine
requirements differ significantly among countries.
In general, developed countries have adopted
standards sooner and have mandated more rapid
phase-in schedules than developing countries.
Canada generally adopts U.S. motor vehicle
standards directly following U.S. implementation.
Thus, similar percentage reductions in BC can be
expected from similar engine categories in Canada.
European and Japanese diesel PM standards have
been reducing PM steadily over the last decade
and are achieving BC reductions similar to those in
the United States. In the next few years, the level
of the standards will be such that DPFs will be used
on almost all new on-road European and Japanese
diesel engines.
In Europe, DPFs were first applied to light-duty
diesels; these requirements are relatively recent, with
the latest standards, known as Euro 5, becoming
effective in 2009. Nonroad diesels will start to
phase in DPFs starting with what are termed Stage
IIIB standards in 2011. The nonroad reductions will
be followed by Euro 6 on-road heavy-duty diesel
standards which will require DPFs on all new trucks
starting in 2013. Likewise some locomotive engines
will have DPFs by 2011 although commercial marine
diesels are not regulated.13
Other countries have adopted or proposed heavy-
duty engine emission standards equivalent to earlier
U.S. or Euro emission standards. In the Americas,
these countries include Argentina, Brazil, Chile,
Mexico, and Peru. In the western Pacific and Asia,
these countries include China, India, the Republic of
Korea, Singapore, and Thailand. China is following
the European emission standard progression
with some time lag; however, China has not yet
implemented low sulfur fuel nationwide to enable
widespread use of DPFs. In Europe outside of the
European Union, Russia and Turkey have adopted
earlier Euro standards. These countries are making
progress in reducing BC emissions from heavy-duty
vehicles. In addressing the future impact of possible
standards, it is important to account for both the
vehicle/engine standards and growth in the number
of vehicles/engines as well as increases in usage
(such as vehicle miles traveled).
13 The European standards use the PMP (particle measurement
program) methodology with thermal treatment (catalytic oxidation)
to remove volatile particles before the PM is measured which
removes much of the organic PM and, thus, PM as measured by the
European test procedure has less organics than that measured by
the U.S. test procedure. With the PMP it is important to distinguish
between particle mass and particle number where organics, which
tend to be small in size, make a contribution. The treatment of
organics is an important distinction for PM control and may affect
the control technology used, which could affect BC reductions.
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Relatively little is known about the costs of DPFs
in other countries. However, it is expected that the
costs for DPFs should not differ greatly from costs
in the United States. More details on diesel PM
emission standards in other countries are discussed
in Appendix 6. It is important to note that few
countries have pursued standards for nonroad
diesels such as construction and farm equipment,
locomotives, and commercial marine vessels
(categories 1 and 2). Such standards, which already
exist in the United States, may offer a mitigation
opportunity internationally.
For control of BC from C3 marine internationally,
EPA is working with the IMO, the Arctic Council,
and others to recommend what can be done to
better define and reduce BC from C3 engines in
international waters. Such work would include
developing a definition for BC emissions from
international shipping. It would also include
considering measurement methods for BC and
identifying the most appropriate method for
measuring BC emissions from international shipping.
It would also include investigating appropriate
control measures to reduce the impact of BC
emissions from international shipping in the Arctic.
Control measures that can be evaluated include
speed reductions, improved routing/logistics,
vessel, propeller and engine modifications, DPFs
(such as non-catalytic ones that could be used with
higher sulfur fuel), water-in-fuel emulsification,
use of slide-valves, and possibly alternative fuels
(MEPC, 2010; UNEP and WMO, 2011a). Some of
these measures have been discussed in a recent
research article (Corbett et al., 2010) and an earlier
Arctic Council report (Arctic Council, 2009). Finally,
the effect of using a distillate diesel fuel (similar to
what is used for diesel trucks and nonroad diesels)
versus HFO on BC emissions should be investigated.
Use of a distillate fuel is expected to result in less
organic emissions and could increase the BC/PM
ratio although the total mass of BC emitted might
decrease.
8.6 Mitigation Approaches for In-use
Mobile Sources in the United States
Though emissions standards for new engines will
reduce emissions over time, existing engines can
remain in use for a long time (20 to 30 years) (U.S.
Census Bureau, 2004). Opportunities to control
BC emissions from in-use vehicles center almost
exclusively on diesel engines. Despite EPA's diesel
engine and fuel standards taking effect over the
next decade for new engines, in-use diesel engines
will continue to emit large amounts of PM and BC,
as well as other pollutants such as NOX, before they
are replaced. For this reason, strategies to reduce
emissions from in-use engines have received a great
deal of attention, particularly because communities
near freight corridors and other large concentrations
of diesel-powered engines are disproportionately
affected by the pollution. EPA's NCDC estimates that
Diesel Emission Reduction Act (DERA) funding could
be used to apply in-use mitigation strategies to 11
million of the on-highway and nonroad engines now
in the U.S. diesel fleet.
A variety of strategies are available to reduce
substantially harmful emissions from in-use vehicles,
and many of these strategies are cost-effective
given the health benefits associated with reducing
PM emissions. As used by EPA, the term diesel
retrofit includes any technology or system that
achieves emissions reductions beyond that required
by the EPA regulations at the time of new engine
certification. Diesel retrofit projects include the
replacement of high-emitting vehicles/equipment
with cleaner vehicles/equipment, repowering or
engine replacement, rebuilding the engine to a
cleaner standard, installation of advanced emissions
control after-treatment technologies such as DPFs,
or the use of a cleaner fuel (U.S. EPA, 2006a).
The BC mitigation potential of diesel retrofits applied
to existing engines depends on several factors,
including engine application (vehicle or equipment
type), engine age, engine size, engine condition
(maintenance) and remaining engine life. One or
more of these factors will dictate the suitability of a
mitigation strategy. Some engines, whether because
of old age, poor maintenance or duty cycle, are not
able to be retrofitted with DPFs. Engines with limited
remaining life or low usage rates are not good
candidates for retrofits when cost-effectiveness is
considered. It can also be technically infeasible to
replace an old engine with a new one in many cases
because of insufficient space in the original vehicle
or piece of equipment. For some of these vehicles,
truck replacement, with scrappage of the original
vehicle, may be the only viable option to reduce
BC emissions. It is also possible for 10%-15% of the
vehicles in a typical fleet to emit 50% or more of
each major exhaust pollutant due to malfunctioning
engine parts (National Academies Press, 2001). This
is one of a variety of important considerations in
developing mitigation strategies.
The NCDC and the SmartWay Transport Partnership
Program are EPA's two primary programs
responsible for reducing emissions from in-use
diesel vehicles and equipment. These programs
support the testing and deployment of numerous
technologies and strategies to reduce emissions
from in-use diesel engines, including BC, and can
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provide immediate reductions. These programs
are described in more detail below, following
a discussion of key retrofit technologies and
approaches for reducing emissions from in-use
vehicles and engines.
8.6.1 Available Retrofit Technologies and
Strategies for In-use Engines
8.6.1.1 Diesel Exhaust After-treatment Devices
Typically, after-treatment diesel retrofit involves
the installation of an emission control device to
remove emissions from the engine exhaust. This
type of retrofit can be very effective at reducing
PM emissions, eliminating up to 99% of BC in
some cases. Of the diesel retrofit devices currently
available, DPFs most effectively reduce BC. For the
sake of completeness, various diesel retrofits are
covered below.14 Further information is available
from NCDC, including a table of emissions reductions
and typical costs for various diesel retrofits.15 EPA
and CARB adhere to rigorous verification processes
to evaluate the performance and reliability of
available retrofit technologies. These processes
evaluate the emission reduction performance of
retrofit technologies, including their durability, and
identify engine operating criteria and conditions that
must exist for these technologies to achieve those
reductions. Federal funding under the NCDC requires
recipients to use EPA or CARB-verified diesel retrofit
technologies for clean diesel projects.
As previously mentioned, DPFs are wall-flow
exhaust after-treatment devices that are effective at
significantly reducing diesel PM emissions by 85%
to 90% and BC emissions by up to 99%. Because
BC exits the engine in solid particle form, DPFs can
reduce BC up to 99%. The small amount of PM
that does penetrate a DPF is composed of mainly
sulfate and OC. DPFs typically use a porous ceramic,
cordierite substrate, or metallic filter to physically
trap PM and remove it from the exhaust stream. The
collected PM is oxidized primarily to CO2 and water
vapor during filter regeneration. Regeneration can
be passive (via a catalyst) or active (via a heat source)
and is necessary to keep the filter from plugging and
rendering the engine inoperative. Regular engine
maintenance is essential to DPF performance.
Passive regeneration occurs when exhaust gas
temperatures are high enough to initiate combustion
of the accumulated PM in the DPF, usually in the
presence of a catalyst, but without added fuel, heat,
or driver action. Active regeneration may require
driver action and/or sources of fuel or heat to
raise the DPF temperature sufficiently to combust
accumulated PM. Active DPFs may be necessary
for lower engine temperature applications, such
as lower speed urban and suburban driving;
otherwise the DPF may become plugged due to an
accumulation of PM.
For large, on-highway trucks, retrofitting passive
DPFs generally costs between $8,000 to $15,000,
including installation, depending on engine size,
filter technology and installation requirements.
Active DPF systems can cost $20,000 for a heavy
duty diesel truck and up to $50,000 for a large piece
of nonroad equipment. Vehicle inspection, data
logging, and backpressure monitoring systems are
required with each installation; these costs along
with installation of the device, are typically included
in the cost of the DPF (U.S. EPA, 2010a). However,
operating costs incurred due to application of
DPFs are not included in the estimates above.
Operating costs could include the differential cost
for using ULSD, fuel economy impacts related to
increased exhaust backpressure, or changes to
maintenance practices related to the use of retrofit
technologies. There is no increased cost for use of
ULSD in the United States because ULSD is now the
predominant diesel fuel used in both highway and
nonroad applications. In addition, data from existing
retrofits show no significant difference in fuel
economy for fleets with and without these retrofit
technologies.16'17
Some diesel retrofit technologies were designed
to reduce other pollutants, such as NOX and
hydrocarbons, and do not significantly impact BC
emissions. Such technologies include:
• Partial Diesel Particulate Filters (PDPFs) provide
moderate (around 30% to 50%) reduction of
PM from diesel exhaust. However, while limited
test data exists on the effectiveness of PDPFs to
reduce BC, it is likely that these devices result in
minimal BC reductions (UNEP and WMO, 2011).
PDPFs typically employ structures to briefly retain
particles for oxidation, structures to promote air
turbulence and particle impaction, and catalysts
to oxidize diesel particles. Partial flow filters
are capable of oxidizing the soluble organic
14 See http://www.meca.org/cs/root/diesel_retrofit_subsite/what_is_
retrofit/whatj.s_retmfit.
15 See http://www.epa.gov/cleandiesel/technologies/retrofits.htrri
16 These cost estimates are from NCDC's Cost-Effectiveness Paper
2006, updated to 2010 dollars.
17 NREL Ralph's Grocery study at: http://www.nrei.gov/docs/
fy03osti/31363.pdf and/or Clean Air Task Force (2009b). The carbon
dioxide-equivalent benefits of reducing black carbon emissions
from U.S. Class 8 trucks using diesel particulate filters: a preliminary
analysis. Available on the Internet at http://www.catf.us/resources/
publications/view/loo.
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fraction of diesel exhaust and likely some BC. As
of October 2010, only three PDPF technologies
were verified by CARB (none by EPA), and these
were only verified for transport refrigeration units
(TRU). These devices cost about $4,000-$8,000
per unit.
• Diesel oxidation catalysts (DOCs) provide minimal
BC reductions. DOCs are exhaust after-treatment
devices that reduce PM, HC and CO emissions
from diesel engines and are widely used as a
retrofit technology because of their simplicity,
relative low cost, and limited maintenance
requirements. DOCs verified by EPA and CARB are
typically effective at reducing PM by 20 to 40%,
though the PM removed by DOCs is composed
largely of OC that comes from unburned fuel and
oil. DOCs are not an effective mitigation strategy
for BC reductions.
• Closed Crankcase Ventilation Systems (CCVS)
provide negligible BC reductions. In many
diesel engines, crankcase emissions or "blow-
by" emissions are released directly into the
atmosphere through the "road draft tube." Closed
Crankcase Ventilation (CCV) devices capture and
return the oil in blow-by gas to the crankcase,
directing HC and toxics to the intake system for
re-combustion instead of emitting them into the
air.
• Selective Catalytic Reduction (SCR) systems inject
a reducing agent such as diesel exhaust fluid
(DEF), a urea solution, into the exhaust stream
where it reacts with a catalyst to reduce NOX
emissions. Most 2010 and newer on-road diesel
engines come equipped with an SCR system and
SCRs are also available as after-treatment retrofits.
SCR systems require periodic refilling of the
reductant and may also be used with a catalyzed
DPF to reduce PM emissions. Coupling engine
design techniques that lead to a reduction of BC
through a low PM engine strategy with a NOX
after-treatment control device such as an SCR has
been an approach used in Europe. SCR systems,
which are effective in reducing NOX by 60 to 80%,
can provide potential BC reductions when the
engine fuel injection timing is changed for lower
PM and higher NOX emissions.
8.6.1.2 Cleaner Engine Strategies
Engine Repower: Significant emissions reductions
can be achieved by repowering, upgrading, or
"reflashing" a diesel engine. Engine repowering (i.e.,
replacing the engine, but not the entire vehicle)
is straightforward, and the benefits are easily
quantified. For example, when an uncontrolled
engine is taken out of service and replaced with a
new engine, the emissions benefits are determined
from the difference in emissions levels of each
engine. The cost of replacing a vehicle or piece of
equipment is much higher than replacing just the
engine. However, not all vehicles/equipment can be
repowered. New engines are not always compatible
with the original vehicle/equipment.
Engine Upgrade: An alternative to vehicle/equipment
replacement and engine repower is "engine
upgrade." An engine upgrade is the process by
which parts of an in-use engine are replaced with
newer components, resulting in lower emissions.
Engine upgrades are normally sold as kits from an
engine manufacturer and include newer mechanical
parts, and, for electronically controlled engines,
changes to the computer program that controls the
engine. Reprogramming the computer that controls
an engine is known as reflash, and it can change
the mix of pollutants in the exhaust stream (e.g., by
changing the injection timing). Engine upgrades,
including "reflashes," are generally less expensive
than replacing an entire engine, but they are only
available for specific engines. Thus, implementation
is limited by the number of upgradable engines
currently in service.
Vehicle/Equipment Replacement: When no
diesel retrofit solutions can be cost-effectively
implemented for a particular vehicle or piece of
equipment, the option exists to retire the vehicle/
equipment from service before the end of its useful
life and replace it with a newer model. While this is
typically the most expensive method of reducing
emissions, this can be the most feasible strategy
for a particular vehicle or piece of equipment. For
example, significant emissions reductions can be
achieved by scrapping older model drayage trucks
at ports and replacing them with newer, clean diesel
trucks. One benefit to replacing an entire vehicle or
piece of equipment is that newer models often have
improved non-engine systems and parts that are
preferred by operators.
8.6.1.3 Other Emissions Reduction Strategies
A variety of other strategies can also reduce
emissions from in-use vehicles. While the precise
impact of such strategies on BC emissions can
be more difficult to quantify than application of
an after-treatment device, these strategies may
substantially reduce emissions, while improving fuel
economy and extending engine life.
Improved Fleet Maintenance Practices: Since a small
percentage of vehicles in a given fleet may be
responsible for a majority of the fleet's emissions,
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Local Retrofit Projects in the United States
Agricultural Vehicle Repowers
The Air Pollution Control District in San Joaquin Valley received $2 million to repower 33 pieces of agricultural
equipment with new engines that meet or exceed EPA'sTier 3 diesel emission standards. Using ARRA funds, EPA
awarded this project because of its long-term economic and immediate health benefits for the community.The
repowered engines are expected to reduce emissions of NOX by over 160 tons and PM by nearly 6 tons.
Locomotive Repower
The Railroad Research Foundation was awarded $2.9 million to repower 4 locomotives that operate as switchers in rail
yards in Baton Rouge, Louisiana.The original locomotives were built with 3,500 horsepower engines in 1985 and 1986,
and the new engines meet or exceed EPA Tier 2 locomotive engine emission standards. Tier 2 locomotive emissions are
one-third those from Tier 0 locomotives.
Shore Power
Massachusetts Port Authority was awarded $400,000 to install shore-side electric power to ships, with a 9-unit shore
connection system serving 18 berths in South Boston. Most vessels dock at the pier 100 to 300 days per year, and
typically run diesel generators for 10 to 14 hours to provide cabin heat, generate power to unload fish, and supply
electricity for other needs. The new on-shore power hook-ups are projected to reduce PM emissions by 96%.
Construction Retrofits
New Jersey Department of Environmental Protection (NJDEP) was awarded $1.73 million to pay for the cost and
installation of pollution control devices on various construction equipment used in New Jersey. Funding under this
program has allowed NJDEP to implement Phase 2 of its existing New Jersey Clean Construction Program to retrofit
non-road equipment used on publicly funded construction projects. The retrofits are projected to reduce PM by 3.8 tons
annually.
one of the first steps for reducing emissions is
to take an inventory and inspect vehicles and
equipment. This information may be used to identify
vehicles in need of repair and find candidates for
other mitigation options. Repair of poorly operating
engines typically decreases emissions and improves
fuel economy. Furthermore, regularly performed
maintenance will extend the life of vehicles and
equipment (Partnership for Clean Fuels and Vehicles,
2009). For example, many manufacturers prescribe
that engines be rebuilt after accumulating a set
number of hours of use. An engine rebuild involves
replacing some old parts and restoring durable parts
to original factory specifications. In some cases, an
after-treatment technology could be installed at the
time of engine rebuild. This would save time since
the vehicle or equipment would not need to be
removed from service any longer than prescribed for
normal maintenance.
Cleaner fuels can lead to BC reductions via multiple
pathways. As previously stated, ULSD fuel is
necessary for diesel particulate filters and other
after-treatment technologies to be effective. Fuel
options such as compressed natural gas (CNG),
liquefied natural gas (LNG), ethanol, and hydrogen
can yield substantial reductions in PM and BC.
However, this requires installation of engines and
fuel systems compatible with these fuels as well as
infrastructure to facilitate storage and delivery of the
fuels. Many U.S. urban fleets of heavy-duty vehicles
have shifted their diesel-fueled vehicles to those
fueled with CNG or LNG. Transit buses and solid
waste collection vehicles are among those fueled
with CNG. Recently, a number of drayage trucks in
Southern California's Port of Los Angeles and Port
of Long Beach have been converted from diesel to
LNG. As previously stated, it will also be important
to determine the effect from increased use of
biofuels such as biodiesel on BC emissions.
Another form of fuel switching is electrification.
As previously stated in this report, power plant
supplied electricity has extremely low emissions
of BC in the United States. If mobile sources can
be powered by electricity, BC emissions can be
reduced. One example of this is cold-ironing (shore
power) at seaports, which allows marine vessels to
shut down their engines and run normal operations
by plugging into electrical connections at docks.
When a vessel is at berth, it typically runs its
auxiliary diesel engines to provide power for normal
operations (referred to as hotelling). For example,
CARB has estimated that 1.8 tons per day of diesel
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PM was emitted by approximately 2,000 hotelling
ocean-going vessels in California in 2006 (Regional
Planning Organization, 2006). Hotelling emissions
can be dramatically reduced if the vessel uses "shore
power" electricity while at port. It should be noted
that emissions of pollutants from other sources
should be considered when pursuing this and other
alternative fuels/energy sources. For example,
electrification shifts the emissions from the mobile
source to the power plant.
Fuel economy improvements may yield reductions in
BC. Some fuel savings devices, such as low-rolling-
resistance tires and aerodynamic technologies (e.g.,
trailer gap reducers, trailer boat tails, and trailer side
skirts) reduce fuel use with little change to engine
operation. These fuel saving devices likely result
in PM reductions; however, additional research
is needed to quantify the emission reductions.
Hybrid vehicles are potential technologies for CO2
reductions, but further research is necessary to
determine the extent of PM or BC reductions.
Idle reduction: Long-duration idling of truck and
locomotive engines consumes an estimated 1 billion
gallons of diesel fuel annually, resulting in thousands
of tons of PM, a significant fraction of which is BC
(i.e., 15-40%) (Gaines et al., 2006; Lim, 2002). It is
important to consider that while BC is a significant
fraction of overall diesel PM, BC/PM ratios differ
during idling. The reduction in PM due to idling has
definite health benefits, and the reduction in fuel
use results in reduced CO2 emissions and, in turn,
climate benefits. However, the net climate benefit
due to reduction in idling PM is less understood.
Furthermore, idling increases fuel and engine
maintenance costs, shortens engine life, increases
driver exposure to air pollution, and creates
elevated noise levels. Idle reduction programs
and technologies are already prevalent in the US.
They serve as one of the simplest and lowest cost
methods to reduce emissions from engines. Because
reducing idling reduces engine operation, emissions
of all pollutants are lower. Strategies for reducing
idling include both operational and technological
methods. Examples of on-board truck technologies
include:
• Automatic engine shut-off devices programmed
to shut down the engine after a preset time limit
• Direct-fired heaters to eliminate idling used to
heat the cab
• Auxiliary power units (APU) or generators to
provide power for cab comfort at rest stops and
eliminate the need to run the truck engine
• Battery or alternatively powered heating and air
conditioning units.
Off-board technologies include truck stop
electrification, which provides conditioned air and
electricity to truck cabs for accessory loads while
at a truck stop. These systems also may provide
telephone, cable TV, and internet access. A majority
of U.S. states and many municipalities have anti-
idling regulation in place to limit idling of vehicles
(American Transportation Research Institute, 2011).
Transportation modal shift: Transportation of certain
goods can be altered to reduce BC emissions and
increase efficiency. Specifically, a shift from trucks
to rail or to sea and inland waterways can reduce
diesel truck PM emissions and alleviate traffic
congestion (Barth and Tadi, 1998; Winebrake et
al., 2008). It is important to note that modal shifts
can result in localized increases in emissions where
goods movement is concentrated, such as ports
and rail yards. While the percentage of BC in total
locomotive PM emissions is roughly equal to that
of diesel trucks (72-78%), diesel engines under idle
or low load, such as occur in intermodal freight
terminals, emit PM with a smaller fraction of BC
(approximately 15-40%). In addition, ship emissions
can exhibit very different characteristics from truck
or locomotive engines, particularly emissions from
slow-speed engines used in ocean-going vessels
(Category 3) burning residual (bunker) fuel. As
described in Chapter 4 on inventories, recent studies
have reported BC to be a minor fraction of PM from
Category 3 marine engines. However, these data are
limited to a few studies. Further research is needed
in order to better characterize ship emissions and to
better understand the effects of modal shifts on BC
emissions.
8.6.2 Cost-Effectiveness of Retrofits
In 2006, EPA published a report on the
cost-effectiveness of heavy-duty diesel engine
retrofits (U.S. EPA, 2006a). The analysis presented
in that report, which was based on data collected
from 2004-2005, estimated the cost-effectiveness
of installing a passive DPF (one that regenerates
removing built-up diesel PM on its own) on a Class
8 truck to be $12,100-$44,100 per ton of PMZ5
reduced. Model year 1994 and newer class 8 trucks
employed in long-haul operation are generally good
candidates for DPFs.
In 2009, EPA published a Report to Congress,
Highlights of the Diesel Emission Reduction Program,
which provides information on the overall cost-
effectiveness of various diesel emissions reduction
strategies funded under the Diesel Emissions
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Reduction Act. The Report estimates that the
average cost-effectiveness of the DERA projects
funded in 2008 ranged from $9,000 to $27,700 per
ton of PM2.5. According to this analysis, which is
currently being updated, diesel retrofit strategies
compare favorably with other emissions reduction
strategies used to attain national ambient air quality
standards that range from $1,000 to $20,000 and as
high as $100,000 per ton of PM2.5 on an annualized
basis. However, most diesel retrofit strategies are less
cost-effective than regulatory programs designed to
set PM emissions standards for new diesel engines,
such as the emissions standards for 2007 and later
model year heavy-duty highway engines.
8.6.3 Applicability of Diesel Retrofits
The ability to install diesel retrofits on different diesel
vehicles and equipment depends on a number of
factors. Not all engine types are equally well suited
to retrofit strategies; for others (e.g., bulldozers),
long engine lifetime may make retrofits the only
feasible option. The on-highway diesel vehicles in
the United States are mostly heavy-duty trucks. The
2002 Census indicated that most trucking companies
are small businesses that own only one to three
trucks. Smaller businesses are less able than large
businesses to absorb capital costs associated with
emissions reductions from diesel engines.
The nonroad engine and vehicle category includes
a diverse range of equipment from lawnmowers
to marine and locomotive engines to construction
machinery. Each category has specific needs and
challenges. Construction equipment, for example,
is often much more expensive with longer useful
life than on-highway vehicles. This adds complexity
when considering mitigation. Vehicle replacement
is difficult for large construction equipment due
to their high costs. In addition, repower options
are only available for certain types of construction
machines due to space limitations in the engine
compartment.
Currently, PM mitigation strategies for marine and
locomotive engines are limited. No DPFs are verified
or certified by federal or state agencies for these
engines. Therefore, upgrading/replacing engines
and fuel switching are currently the two most
viable mitigation strategies for these engines. Fuel
switching could also include the use of shore power
for larger marine vessels, which eliminates local PM
emissions while ships are at port. New emissions
reduction technologies are being developed to
reduce locomotive and marine emissions. For
example, marine engine upgrade kits have been
implemented with funding support from the EPA
Emerging Technologies Program.18
8.6.4 Experience with Diesel Emissions
Reduction Programs in the United States
Federal, State, and local agencies have
demonstrated substantial capacity to develop and
implement diesel emissions reduction programs.
Collectively, these agencies, in partnership with
environmental and industry stakeholders, have built
a strong foundation for the testing, verification and
implementation of new technologies and strategies.
Many of these programs provide funding or other
incentives for voluntary diesel retrofits, engine
replacements, or idle reductions. These programs
include EPA's NCDC and the SmartWay Transport
Partnership; FHWA's Congestion Mitigation and Air
Quality (CMAQ) Improvement Program; the Texas
Emissions Reduction Plan (TERP), and California's
Carl Moyer Memorial Air Quality Standards
Attainment Program.
8.6.4.1 National Clean Diesel Campaign (NCDC)
The National Clean Diesel Campaign (NCDC)
is a partnership that aims to accelerate the
implementation of emissions control strategies
in the existing fleet through approaches such
as retrofitting, repairing, replacing, repowering,
and scrappage of diesel vehicles and equipment;
reducing idling; and switching to cleaner fuels. This
ten-year effort by EPA to bring together industry,
environmental groups, local and State governments
and Federal programs has resulted in significant
experience with various fleet types and technologies
and reduced emissions from thousands of engines.
Several initiatives through the Campaign have
targeted specific sectors, such as Clean School Bus
USA and the clean ports program, demonstrating
a variety of technologies and strategies on those
fleets.
In 2005, a dedicated source of funding was
authorized by Congress for implementation of
NCDC projects on a wider scale. The Energy Policy
Act of 200519 provided EPA with grant and loan
authority to promote diesel emissions reductions
from the existing in-use fleet in the United States
and authorized appropriations of up to $200 million
per year to the Agency under the DERA provisions
for FY2007 through FY2011. The DERA Program may
serve as one of the best avenues and foundations
for reducing BC emissions in the United States (U.S.
18 See http://www.epa.gov/cleandiesel/projects/projects.htm.
19 http://www.gpo.gov/fdsys/pkg/PLAW~109publ58/pdf/PLAW~
109publ58.pdf
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Chapter 8
EPA, 2009a). Congress appropriated $169.2 million
in funding under this statute in FY 2008 through
FY 2010. In addition, the American Recovery and
Reinvestment Act of 2009 allotted the NCDC $300
million. The Diesel Emissions Reduction Act of
2010 was signed into law in January 2011. This law
authorizes DERA for $100 million per year from
FY2012 through FY2016.
DERA offers a funding vehicle for immediate BC
reductions within the in-use fleet. The first year of
DERA funding reduced emissions from more than
14,000 diesel-powered highway vehicles and pieces
of nonroad equipment. DERA funding supported
a wide range of verified technologies, cleaner
fuels, and certified engine configurations, such as
repowers, replacements, idle-reduction technologies,
biodiesel, and retrofit devices such as DPFs. DERA
funding also supported diesel programs in state
governments.
The diesel emissions reductions resulting from the
FY2008 grants for PM will total approximately 2,200
tons by 2031, which translates to 1,540 tons of BC
reductions, assuming 70% of PM is BC. The health
benefits will range from a net present value of $580
million to $1.4 billion, including an estimated 95 to
240 avoided premature deaths.
From 2008-2010, EPA received applications
requesting more than $665 million, which equates
to $7 for every $1 available in clean diesel funding.
Thus, there remains strong interest in utilizing DERA
to reduce diesel emissions. Additionally, a large
number of high emitting engines remain currently
in use. In moving forward with the program, a
few challenges remain. For example, there are too
few verified technologies for nonroad and marine
engines and older diesel trucks, limiting the extent of
achievable emissions reductions. The nonroad sector
offers another challenge because of the number
and diversity of nonroad engine types, the range of
horsepower and the varying usage and duty cycles
of the equipment.
Because BC is a regional pollutant, EPA, through
the DERA Program, provides assistance to state and
local governments in developing their own clean
diesel programs. This includes targeting current
nonattainment areas where clean diesel strategies
can assist in meeting local emissions reduction
goals and providing high quality data to states that
depend on the performance of diesel emissions
reduction strategies in their air quality plans. In
addition, EPA conducts in-use testing—confirming
the performance of verified technologies in the
field—while working cooperatively with industry
groups, engine manufacturers, and state agencies to
expand the list of clean diesel technology options.
8.6.4.2 SmartWay
In 2004 EPA launched its SmartWay Transport
Partnership. SmartWay is an innovative, voluntary
partnership between EPA and private industry
to reduce fuel use and emissions from goods
movement. SmartWay promotes fuel-saving
technologies and emission control technologies;
some technologies—like idle reduction or newer
truck replacements—do both. Since most cargo-
hauling large trucks, locomotives, barges, and other
freight vehicles use diesel fuel, and these vehicles
remain in the legacy fleet for decades, reducing
fuel use and emissions from goods movement and
the legacy fleet can have a major impact on diesel
emissions, including emissions of BC.
SmartWay provides shippers as well as truck
carriers with standardized tools and approaches
to assess, benchmark, track and reduce fuel use
and emissions from goods movement. SmartWay
offers technical assistance to enable partners
to improve performance. The program offers
incentives (SmartWay logo eligibility, SmartWay
partner ranking, recognition of top performers)
to encourage continual improvement. To enable
this improvement, SmartWay helps its shipper and
carrier partners identify fuel-saving operational and
technical solutions through its technology program.
This technology program researches and evaluates
fuel-saving technologies, develops standardized
protocols for the measurement of technology
improvements (e.g., fuel consumption, aerodynamic
impacts, long-duration idle reduction), and officially
verifies the benefits of certain technology types
(i.e., long-duration idle reduction technologies,
low rolling resistance tires, and aerodynamic
components).
While a wide variety of technologies exist to
reduce fuel consumption and costs for trucking
companies, many companies lack the up-front
investment capital to benefit from them. The
SmartWay Finance program, funded by diesel
emissions reduction funding, aims to accelerate
the deployment of energy efficiency and emissions
control technologies by helping vehicle/equipment
owners overcome financial obstacles. Since 2008,
the SmartWay Finance program has awarded over
$30 million to help small trucking companies reduce
fuel costs and emissions. These innovative loans
help small trucking firms reduce PM emissions, and
lower their fuel costs by purchasing newer used
trucks equipped with idling and emissions reduction
technologies.
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Nearly 3,000 companies, from small firms to Fortune
500 companies, belong to SmartWay. To date, these
SmartWay partners have saved $6.1 billion dollars
by cutting their fuel use by 50 million barrels of
oil. This is equivalent to taking 3 million cars off
the road for an entire year. Improving supply chain
efficiency helps these companies grow the economy,
protect and generate jobs, cut imports of foreign
oil, contribute to energy security, and be good
environmental stewards.
In developing new national standards to bring
cleaner, more efficient trucks to market, EPA and the
Department of Transportation's National Highway
Traffic Safety Administration (NHTSA) drew from
the SmartWay experience. This experience includes
developing test procedures to evaluate trucks and
truck components and determining how these
features and components perform. While focused
on American freight-efficiency, SmartWay has
responded to industry demand to recognize the
importance of the global supply chain by expanding
its tools and building the capacity for SmartWay-
based programs in other countries.
8.6.4.3 Congestion Mitigation and Air Quality
Improvement Program (U.S. DOT)
The Congestion Mitigation and Air Quality (CMAQ)
Improvement Program, jointly administered by
the U.S. Department of Transportation's Federal
Highway Administration (FHWA) and the Federal
Transit Administration (FTA), provides roughly
$1.7 billion in annual funding for a variety of
emissions reduction projects including transit, traffic
signalization, bicycle/pedestrian facilities, demand
management, and diesel retrofit projects. According
to the most recent data available, between 2005 and
2007, approximately $285 million of CMAQ funds
were spent on diesel retrofits. New priority for the
funding of diesel retrofit projects was established
by Congress with the Safe, Accountable, Flexible,
Efficient Transportation Equity Act: A Legacy for
Users (SAFETEA-LU) in 2005.
The allocation of CMAQ funds is managed by the
state DOTs. CMAQ aims to implement projects that
will help areas attain or maintain the NAAQS. Diesel
retrofits are more cost-effective in reducing PM than
other typical CMAQ projects, such as traffic signal
optimization (Diesel Technology Forum, 2006, 2007).
8.6.4.4 State Programs
Mandatory retrofits: The state of California has
enacted legislation to require in-use heavy duty
diesel fleets to meet minimum emission standards.
The legislation is implemented through CARB and
applies to many sectors, including both on-highway
and nonroad diesel engines. Most of the regulations
require accelerated fleet turnover, which includes
repowering or retiring vehicles, or requiring best
available control technology (BACT) to be installed
on diesel engines. Almost all on-highway heavy-
duty diesel vehicles, including buses, drayage trucks,
and class VIII trucks will be required to reduce diesel
emissions.
Several states have passed legislation similar to
California's. New Jersey has instituted a mandatory
retrofit program requiring owners of diesel vehicles
to retrofit with best available retrofit technology
(BART). The state reimburses vehicle owners/
operators for all expenses. New York has also
instituted a mandatory retrofit program that applies
to all heavy-duty state-owned and contractor
vehicles.
Incentive programs: The state of California's Carl
Moyer Memorial Air Quality Standards Attainment
Program provides incentive grants for cleaner-than-
required engines, equipment, and other sources
of pollution providing early or extra emission
reductions. The program started in 1998 and has
funded hundreds of millions of dollars worth of
projects since its inception. California voters also
passed Proposition IB in 2006, which allocated
$1 billion to reduce air pollutant emissions from
freight along California's trade corridors. Both of
these incentive funding programs rank applicants
based on cost effectiveness (e.g., $/ton). Carl Moyer
funds cannot be used to fund compliance with state
or federal laws. Thus, funding opportunities are
becoming limited due to California's implementation
of regulations affecting most categories of mobile
sources.
The Texas Emissions Reduction Plan (TERP), a
program of the Texas Commission on Environmental
Quality (TCEQ) provides financial incentives to
eligible individuals, businesses or local governments
to reduce emissions from polluting vehicles and
equipment in the state of Texas. TERP has provided
over $797 million since 2002, affecting over 12,500
diesel engines with engine/vehicle replacement as
one of the key clean diesel strategies. Though this
incentive program focuses more heavily on NOX,
there is still an opportunity for manufacturers to
develop both NOX and PM combination technology
strategies for BC reductions, through the New
Technology Research and Development Program
(NTRD), which encourages and supports the
research, development, and commercialization of
technologies that reduce pollution in Texas.20
' http://www. fceq. fexas.gov/airquality/ferp.
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Chapter 8
8.7 Mitigation Approaches for In-use
Mobile Engines Internationally
There are millions of large diesel-powered vehicles
throughout the world, including buses, heavy duty
trucks, off-road vehicles, locomotives, and marine
vessels. The exact size of the international diesel
fleet is not easily characterized. Some countries
are similar to the United States in one or more of
the following: vehicle registration, inspection and
maintenance programs, availability of low-sulfur
fuel, technology certification/verification programs,
and readily available technologies. However, many
(mainly developing) nations have little to none of
this infrastructure in place. Furthermore, developing
countries tend to have older and less well-
maintained engines and vehicles than developed
countries, and the availability of low-sulfur diesel fuel
is limited. Therefore, many engines in developing
countries are not good candidates for tailpipe
control strategies like passive DPFs. In addition,
the costs of DPFs may be prohibitive for some
countries. Most retrofit programs around the world
(including in the United States) have relied heavily
on government funding, which presents a significant
financial challenge.
EPA has often advised other nations and supported
international demonstration projects in an effort
to transfer information and technologies to those
that seek to reduce emissions from mobile sources.
Additionally, EPA's diesel retrofit experts have
advised and participated in several pilot retrofit
projects where diesel trucks and buses were fitted
with various exhaust after-treatment devices. Low-
sulfur diesel was obtained for the projects in most
cases. The projects have shown generally that, if
appropriate fuel is provided and engine maintenance
is addressed, DPFs are viable options to reduce PM
(and thus BC) on some vehicles. Following a relatively
small EPA supported pilot project in Beijing in
2006, city authorities went on to retrofit more than
6,000 vehicles with active DPFs prior to the Beijing
Olympics. That number is now above 8,000 and
growing. EPA has also assisted in retrofit projects in
Mexico City, Bangkok, Santiago, and Pune (India).
As noted earlier, the SmartWay program recognizes
the importance of the global supply chain and
has shared its program and technology expertise
with other countries. EPA hosted a SmartWay
International Summit in December of 2008 to offer
guidance to numerous countries which are also
developing freight sustainability programs. As a
result of that Summit and other capacity building
and information sharing, multiple countries and
regions have gone on to implement SmartWay-like
programs. Mexico, Canada, France and Australia
have each developed and launched freight
sustainability programs using SmartWay templates
for tools and program design, partnership structures
and best practices. Additionally, a consortium of
SmartWay Partners and other businesses in the
European Union have beta-tested SmartWay tools
with the intent of developing a SmartWay platform
for implementation throughout the region.
More recently, EPA has collaborated to help China
develop multiple freight sustainability projects
utilizing SmartWay technologies and program
design elements. EPA first provided technical
expertise for the Green Truck Pilot in Guangzhou
in 2009. The World Bank funded the retrofitting
of SmartWay technology on local trucks and
demonstrated notable fuel savings. Based on those
results, the World Bank secured funding from the
Global Environmental Facility for the Guangdong
Green Freight Demonstration Program. This $18
million project will implement truck retrofits
and upgrades using SmartWay technologies
and financing methods, as well as logistical
improvements, driver training and capacity building
for governmental officials. The Ministry of Transport
and the China Sustainable Energy Program are
developing a nationwide freight efficiency program,
built in part on these pilot projects.
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Chapter 9
Mitigation Approaches for
Stationary Sources
9.1 Summary of Key Messages
• There has been a dramatic decline in BC emissions
from industry in developed countries over the last
century. Stationary sources in the United States
now account for only 8% of the U.S. BC inventory;
sources include industrial, commercial, and
institutional (ICI) boilers; power plants; industrial
processes such as cement manufacturing; and
stationary diesel engines.
• Internationally, emissions from stationary sources
account for about 20% of the global inventory,
with highest emissions in China, the former USSR,
India, and central/South America. Main sources
are brick kilns, coke ovens (largely from iron/steel
production), and industrial boilers.
• Available control technologies and strategies
include direct PM2.5 reduction technologies
such as fabric filters (baghouses), electrostatic
precipitators (ESPs), and diesel particulate filters
(DPFs). Once installed, these strategies range in
cost-effectiveness from as little as $48/ton PM2.5
to $685/ton PM2.5 (2010$) or more, depending
on the source category. However, they also may
involve tens of millions in initial capital costs.
Additional source testing data is needed to clarify
the efficiency of these controls for removing BC
specifically.
• Internationally, emissions from a number
of source categories may grow as countries
industrialize. Reducing emissions from smaller,
inefficient facilities may require phasing out or
replacing the entire unit, while larger facilities can
apply many of the existing PM filter technologies
already in commercial use. However, both of these
options may be associated with substantial cost
and implementation difficulties.
9.2 Introduction
Emissions of BC from stationary sources1 generally
represent a smaller portion of current global
inventories than mobile sources and other source
categories. As mentioned in Chapter 4, this is due
in large part to a significant decline in industrial
BC emissions from developed countries over the
past century. These reductions have been achieved
through improved combustion, shifts in fuel use,
and application of control technologies to limit
direct PM emissions. Although some uncertainty
remains regarding the exact efficiency of these
control techniques for reducing the BC fraction
of PM2.5 emissions, that uncertainty does not
change the conclusion that emissions of BC from
U.S. stationary sources are relatively modest in
comparison to other key sectors of the national
inventory. In contrast, stationary sources represent
a larger fraction of international inventories, and in
some regions these sources are key contributors to
overall direct PM2.5 emissions which adversely affect
public health and the environment. Thus, continued
mitigation of stationary source BC emissions
domestically and internationally will lead to
improved public health and will also provide climate
co-benefits.
There are a number of relatively well-developed
control technologies that have successfully been
applied to reduce direct PM25 (including BC) from
stationary sources. This section discusses PM control
technologies and strategies that are applicable
to BC mitigation from domestic and international
stationary sources. Where possible, it provides
information about the applicability, performance,
and costs of these approaches. Since these control
technologies are well-established, much of this
information is drawn from EPA and other control
technology guidance documents developed for PM
mitigation purposes.
1 The term "stationary sources" as used in this chapter refers to
large and small industrial or combustion operations. It does not
include residential fuel combustion for heating or cooking.
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Chapter 9
9.3 Emissions from Key Stationary
Source Categories
The combustion of fossil fuels such as coal or oil
is often the primary source of BC emissions at an
industrial facility. In the United States and other
developed countries, stationary source emissions
of BC have been reduced substantially from historic
levels. As discussed in Chapter 4, current emissions
from stationary sources (including both "industrial
sources" and "fossil fuel combustion" categories in
the U.S. inventory) account for roughly 8% of the U.S.
BC inventory (see Table 4-2). These emissions come
from industrial, commercial, and institutional (ICI)
boilers; power plants; and other types of industrial
sources, such as cement manufacturing or stationary
diesel engines used for many purposes including
irrigation or oil and gas extraction.
Stationary sources account for a slightly higher
percentage (20%) of total worldwide BC emissions,
and more than 30% of BC emissions from contained
combustion (i.e., sources other than open biomass
burning) (see Table 4-6 and Bond et al., 2004). In
certain developing world regions, such as China and
India, stationary sources represent a very significant
percentage of the BC inventory. The regions with the
highest percentage of "contained" BC emitted from
industry and power generation are China, the former
USSR, India, and central/South America (Zhang et
al., 2007). Key source categories include brick kilns,
coke production/iron and steel production, and
industrial boilers. As discussed in Chapter 7, however,
BC emissions from industrial sources are expected to
decline worldwide under most scenarios. This decline
is anticipated to occur in developing countries as well
as developed countries.
In the United States, direct emissions of PM and
BC from stationary sources have been reduced
significantly due to improved combustion efficiencies
in industrial operations and implementation of
federal and state clean air regulations over the past
several decades. This declining emissions trend is
expected to continue as further reductions will be
needed to meet revised air quality standards and
mitigate adverse effects on public health and the
environment. EPA's modeled emissions inventory
projections indicate that direct PM emissions from
stationary sources are expected to decline by
about 20% between 2005 and 2020. For example,
sources in nonattainment areas will be required to
implement emissions reduction strategies to help
areas attain the 1997, 2006, and any future revisions
to the PM2.5 NAAQS. Certain facilities will also be
required to comply with revisions to maximum
achievable control technology (MACT) and new
source performance standards (NSPS) for specific
source categories. These standards will lead to
control of some sources that currently do not have
any PM controls; they will also lead to improved
levels of control for certain sources that already
have PM controls. Older power generation sources
may be retired as well. However, in an overall sense,
near-term BC emissions reductions from domestic
stationary sources are expected to be modest when
compared to expected reductions in other sectors,
such as the mobile source category.
In general, stationary sources burning coal dominate
the U.S. BC emissions inventory for stationary
sources. However, many of these sources have
high combustion efficiencies and have already
applied substantial emissions controls. For example,
nearly all large coal-fired EGUs have electrostatic
precipitators (ESPs) or fabric filters for PM control.
Estimates by the U.S. Department of Energy indicate
that 76% of fossil-fuel steam-electric generating
units in the United States (1,194 of 1,568) have some
form of PM control—and those that do not are likely
to be fueled by natural gas (USEIA, 2010). More than
80% of these sources operate ESPs, while about 14%
have fabric filters. These control technologies are
described further, below.
ICI boilers are a wide-ranging category of
combustion units that collectively can burn a wide
variety of combustible fuels, including coal, oil,
natural gas, and biomass. There are thousands of
ICI boilers across the country, varying in size from a
few million Btu/hr for small commercial or industrial
units to over 10 million Btu/hr for large boilers.
Their operations range from intermittent to near-
steady state. Most large units are covered under
new regulations issued in April 2011 that include
stringent standards for PM, mercury, and certain
hazardous air pollutants, although EPA is currently
reconsidering these standards.2 Under the proposed
reconsideration of this rule issued in December
2011, EPA has projected that the new emissions
limits applicable to major source boilers and process
heaters would reduce PM2.5 emissions from these
sources by 41,200 tons by 2014.
Stationary engines burning diesel fuels also account
for substantial BC emissions. These engines are
similar to mobile diesel engines and typically use
the same fuels, but they can also operate using
natural gas or heavier fuel oil grades than mobile
diesel engines. They are used to perform a range
of different tasks, such as pumping water or oil
2 The April 2011 final rule is available at http://www.epa.gov/
airquality/combustion/actions.html#febll. EPA issued a proposed
reconsideration of this rule on December 2, 2011 (see http://www.
epa.gov/airquality/combustion/actions.html), and expects to issue a
revised final rule in 2012.
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through pipelines, operating equipment in remote
locations, or providing backup power generation.
Many other categories of industrial sources emit
relatively low amounts of BC. In the current U.S.
inventory, the "natural gas combustion" sector
appears to have substantial BC emissions, but
this is likely due to severe constraints on the data
used to generate these estimates.3 Given current
knowledge of the utility and major source boiler
inventory and the mechanisms of BC formation,
EPA does not believe that there are significant BC
emissions from natural gas combustion sources with
good combustion practices. It is recommended that
additional source testing and research be conducted
to improve current emission factors associated with
natural gas combustion. It is also recommended that
additional source testing and research be conducted
on the related category of oil and gas flaring (see
additional discussion in section 9.7.4 below).
Another category of note is use of biomass for power
and steam generation. While wood-fired boilers are
currently a fairly small part of the U.S. inventory,
there is the possibility that more stationary sources
may increase their use of biomass as a fuel source
with the intention of reducing their carbon footprint.
To the extent that sustainable biomass becomes a
more common source of fuel, BC emissions could
rise in absolute terms if not effectively controlled.
Fortunately, effective technologies are already
available on the market that can control emissions
from these sources, as described below.
9.4 Available Control Technologies for
Stationary Sources
This section provides an overview of the main
technologies for reducing PM2.5 emissions from
stationary sources. Several post-combustion PM
control technologies have been in operation for
many years and have been demonstrated to be
quite effective in reducing PM2.5. These technologies
are also considered to be relatively effective at
controlling BC because BC is a filterable component
of PM2.5. Many studies to date have assumed that
PM2.5 control technologies will reduce similar
fractions of PM25 and BC mass. However, it has also
been recognized that reduction efficiency declines
to some extent as particle size decreases (and BC
particles are commonly smaller than 1 micrometer in
3 The current AP-42 emissions factor for BC from natural gas
combustion is considered to be highly questionable. Bond et al.
(2006b) indicated significantly lower emissions factors for industrial
natural gas combustion than that published in AP-42. Bond reported
an emission factor of 0.004±0.004 g PM per kg fuel, two orders of
magnitude lower than the 0.21 g/kg found in AP-42.
diameter). For this reason, it is recommended that
additional source testing and research be conducted
on stationary sources to better understand control
efficiencies for BC and to develop improved
emission factors for specific source categories.
The two most effective control technologies
for PM2.5 (and therefore for BC) are fabric filters
(sometimes called baghouses) and ESPs. Although
there are other technologies used to reduce
emissions of PM (such as cyclones and Venturi
scrubbers), they are often designed to control
larger particles (PM10 and larger), and therefore
are considered to be less effective in terms of BC
mitigation. EPA provides a thorough overview
of the principles of operation, design variations,
applicability, performance, and associated costs of
fabric filters and ESPs in the 2002 EPA Air Pollution
Control Cost Manual (see U.S. EPA, 2002b).
9.4.1 Fabric Filters
A fabric filter unit consists of one or more isolated
compartments containing rows of fabric bags in
the form of round, flat, or shaped tubes, or pleated
cartridges. Particle-laden gas passes up (usually)
along the surface of the bags then radially through
the fabric. Particles are retained on the upstream
face of the bags, and the cleaned gas stream is
vented to the atmosphere. The filter is operated
cyclically, alternating between relatively long periods
of filtering and short periods of cleaning. During
cleaning, dust that has accumulated on the bags is
removed from the fabric surface and deposited in
a hopper for subsequent disposal. Fabric filters are
not recommended for boilers burning oil because
particles from oil combustion are sticky and tend to
clog the filter.
A properly designed and well run baghouse will
generally have extremely high particle collection
efficiencies (i.e., greater than 99.9%). Baghouses
are particularly effective for collecting fine particles
from power generation and a range of industrial
facilities. For example, tests of bag houses on two
utility boilers (Broadway and Cass, 1975; Cass and
Broadway, 1976) showed efficiencies of 99.8 % for
particles 10 u.m in diameter and larger and 99.6% to
99.9% for particles 2.5 u.m in diameter and smaller.
Studies have shown that collection efficiencies
greater than 99% can be achieved for particles less
than 1 u.m in diameter (NESCAUM, 2005; Buonicore
and W.T. Davis (eds.), 1992). A recent report for the
U.S. Forest Service on the applicability of different
PM emissions control technologies to small wood-
fired boilers found that mechanical collectors, such
as multicyclones, were only modestly effective in
reducing PM emissions, with an average of about
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Chapter 9
15% control efficiency. In the Forest Service study,
fabric filters achieved 74% reduction of PM2.5, even
with some of the uncontrolled flue gas circumventing
the baghouse (Hinckley and Doshi, 2010).
9.4.2 Electrostatic Precipitators
An ESP is a particle control device that uses an
electrical charge to move the particles out of the
flowing gas stream and onto collector plates.
Appropriately designed ESPs are effective at
removing particles from sources operating at high
temperatures and having large volumes of gas.
They operate on many types of facilities that emit
PM and BC. ESPs typically achieve greater than
99% PM removal efficiency (depending upon the
design parameters of the specific unit), although
the removal efficiency is generally lower for
submicrometer size particles such as BC. Smaller
particles are more easily carried by the gas stream,
and therefore the ESP collection efficiency for very
fine particles like BC is typically lower than the
efficiency for removing larger particles (i.e. greater
than 1 micrometer in diameter). In general, fabric
filters are more effective than ESPs at removing BC.
To address the lower ESP collection efficiency on
submicrometer particles, a hybrid PM collection
system can be employed. Some designs place the
baghouse downstream of an existing ESP to improve
overall collection efficiency. Others integrate the ESP
and baghouse components. This type of system can
achieve 99.99% control of all particle sizes from 0.01
to 50 micrometers (Zhu, 2003).
Sources such as biomass combustors that also
generate significant levels of condensable PM may
benefit from wet ESP designs, in which the collector
surfaces are washed with water (either continuously
or intermittently) to clean the particles from the
collectors. Tubular ESPs are most commonly used for
operations where the PM is either wet or sticky (U.S.
EPA, 2002b, Chapter 1). Typical applications include
sulfuric acid plants, coke oven by-product gas
cleaning (tar removal), and recently in iron and steel
sinter plants. Because wood combustion systems
in particular can produce PM that is sticky, tubular
ESPs may be appropriate for use in small systems for
reduction of PM and BC.
A relatively new technology known as an
agglomerator can also be used in conjunction with a
control device (such as an ESP) in utility or industrial
applications. This technology is installed in the high
velocity ductwork leading to the control device. It
pre-treats the dust particles prior to entering the
device, agglomerating small and large particles
together, thereby making it easier for the control
device to collect the larger particles. It has been
shown to improve the ESP collection efficiency of
very fine particles (less than 1 micrometer in size)
by 75-90% (Truce and Wilkison, 2008).4 There are
a number of commercial installations of the Indigo
Agglomerator technology in place; most installations
are upstream of an ESP, but this technology has also
been successfully operated in conjunction with a
fabric filter or wet scrubber.
Wet particulate scrubbers are generally not
appropriate for control of BC. Collection efficiencies
for wet scrubbers vary with the particle size
distribution of the waste gas stream. In general,
collection efficiency for submicrometer particles is
much lower than for ESPs or fabric filter systems.
Submicrometer particle collection efficiencies for
wet scrubbers typically are on the order of 50% or
less, although cyclonic wet scrubbers may be able to
remove as much as 75% of submicrometer particle
mass (U.S. EPA, 2002b).
9.4.3 Diesel Particulate Filters and
Oxidation Catalysts
There are more than a million stationary diesel
engines in use today and together these sources
have substantial emissions of PM and NOX. For most
diesel engines, BC is a significant component of
untreated exhaust; these emissions can be reduced
through DPF technology.
DPFs were originally developed for mobile engine
applications (see Chapter 8). They include variations
such as diesel particle traps and catalytic and
noncatalytic soot oxidation systems. These units
typically involve mechanical filtering of soot
particles and a mechanism for oxidation of the soot
to CO2. This second step is sometimes referred
to as regeneration, and eliminates the need for
collecting and disposing of the captured particles.
Catalysts are used to enhance the oxidation process.
Depending upon the design and operation of the
DPFs, removal efficiencies of between 40% and 99%
can be achieved (van Setten et al., 2001).
To ensure optimal performance of DPFs and to
avoid poisoning of the catalyst, the diesel engine
should burn fuel with low sulfur content. DPFs have
been identified by the California Air Resources
Board (CARB) as a verified technology for stationary
engines serving prime and emergency standby
generators and pumps.5 In some situations, such
as where loads are not transient and exhaust
4 See also http://www.indigotechnologies.com.au/agg_overview.php.
5 A summary of CARB-verified diesel emission control strategies is
located at http://www.arb.ca.gov/diesel/verdev/vt/cvt.htm.
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temperatures are high enough, non-catalytic DPFs
can be used with fuels having a higher sulfur content.
EPA issued new source performance standards
in July 2006 (71 FR 39153) for new compression
ignition (CI) stationary internal combustion engines.
These standards implemented new restrictions on
emissions of PM, NOX, VOC, and CO as well as new
limits on the level of sulfur permitted in diesel fuel. In
June 2011, EPA further strengthened these standards,
establishing stringent new performance standards
for stationary CI engines with a displacement of 10
to 30 liters per cylinder (76 FR 37954), consistent
with recent revisions to standards for similar mobile
source marine engines.6
9.5 Cost-Effectiveness of PM Control
Technologies
The cost-effectiveness of an emissions control device
for a stationary source is often expressed in terms
of dollars per ton of pollutant reduced. Factored
into this amount are capital costs (amortized over
several years) for design and installation of the
control equipment, and annual costs for operating
and maintaining the equipment. Because many
emission standards for stationary sources to date
have included emission limits for total filterable
PM (as opposed to PM2.5 or BC), many of the cost-
effectiveness values found in published reports are
expressed in terms of the cost per ton of reducing
total PM. It has been noted earlier that we assume
that BC emissions will be reduced with the operation
of a fabric filter or ESP on a stationary source.
However, it is acknowledged that actual control
efficiencies for capturing submicrometer BC particles
are uncertain, and they are likely to be somewhat
lower than the assumed control efficiencies for total
PM or PM2.5. For this reason, additional research
and source testing is needed to develop improved
measurement techniques and development of robust
emission factors for specific source categories.
The effectiveness of a given control technology used
for a specific source category will depend not only
upon the performance of the particular technology,
but also upon the level of control that is already
in place. For instance, most large coal combustion
sources, such as EGUs, are likely to be well controlled
to comply with prior PM emission standards. In
contrast, smaller and older coal combustion units
that have not been subject to similar emission
standards may in some cases be able to reduce PM
emissions for a lower dollar cost per ton because
installation of the same technology will remove
a greater mass of PM (including PM25 and BC)
compared to a well-controlled EGU. Therefore,
some sources that have been completely exempt
from PM control because of their age, small size,
or limited operation (such as certain distillate oil or
coal combustion systems) may present favorable
mitigation opportunities. Thus, a reasonable and
cost-effective mitigation strategy requires detailed
knowledge of the sources and their emissions,
on both a per-source basis and across the full
population of those sources.
The 2002 EPA Air Pollution Control Cost Manual
and several related 2003 control technology fact
sheets provide typical cost effectiveness ranges for
PM reduction by fabric filters and ESPs. The cost-
effectiveness range identified for a fabric filter was
$51 to $462 per ton (2010$); for an ESP it was $48 to
$685 per ton (2010$).
Table 9-1 presents information on PM control
cost ranges as adapted from multiple sources
referenced in a 2009 NESCAUM report for ICI
boilers (NESCAUM, 2009). Capital costs can vary
significantly depending on the source specific
characteristics. For some large utility boilers (500
megawatts), a fabric filter can require an investment
on the order of $70 to 105 million (2010$)
(NESCAUM, 2010). Cost-effectiveness values per ton
are commonly higher for oil combustion units than
for those burning coal or wood. In general, however,
PM control technologies are well-established.7
The reduction of BC and PM2.5 from industrial
categories can provide significant public health
benefits, particularly for communities located
close to emissions sources. For example, recent
regulations on different sizes of ICI boilers in the
U.S. were estimated to provide $73,000 - $294,000
in health benefits per ton PM2.5 reduced (2010$)
(U.S. EPA, 2011a).8 This suggests that controlling BC
6 In addition, the action revises the requirements for engines with
displacement at or above 30 liters per cylinder to align more closely
with recent standards for similar mobile source marine engines, and
for engines in remote portions of Alaska that are not accessible by
the Federal Aid Highway System. See http://www.epa.gov/ttn/atw/
nsps/sinsps/fr28jnl 1 .pdf.
7 It should also be noted that control of PM and BC results in
millions of pounds of particulates being captured and disposed
of as solid waste; and in some cases it is discharged as part of
wastewater discharges. The problems associated with disposal
of coal combustion residues are well recognized, and it will be
important to manage such wastes effectively in the future.
8 For major boilers: $73,000 to $182,000 per ton of directly emitted
fine particles (range from Pope to Laden, 3% discount rate, 2014
analysis year, 2010$). For area boilers: $122,000 to $294,000 per
ton of directly emitted fine particles (range from Pope to Laden, 3%
discount rate, 2014 analysis year, 2010$). See http://www.epa.gov/
ttn/ecas/reqdata/RIAs/boilersriafinalll0221_psg.pdf, Tables 7-2 and
7-3.
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Chapter 9
Table 9-1. PM Control Costs for ICI Boilers. (Source: U.S. EPA, based on data from (NESCAUM, 2009). Costs
have been updated to 2010$.)
r»/i r. j *• Size of
Fuel Type Technology PM Reduction
Potential (mmBTU/hr)
Capital
Costs
(Million,
2010$)
Cost Effectiveness ($/
ton removed, 2010$)
Reference
Coal
Coal
Oil
Wood
Wood
Dry ESP
Fabric Filter
Dry ESP
ESP
Fabric Filter
90-99%
90-99%
90-99%
99.5%
99.5%
250
250
250
Medium
Medium
$3.5-47
$2.4 - 27
$2.4 - 26
—
—
$184-1,529
$498-1,183
$2,739-24,715
$239-344
$173-293
MACTEC (2005)
MACTEC (2005)
MACTEC (2005)
STAPPA (2006)
STAPPA (2006)
emissions from industrial sources may be highly
cost-effective and should remain a part of any overall
BC reduction strategy.
9.6 Mitigation Approaches Other than
PM Control Technologies
9.6.1 Process Modification/Optimization
As a product of the combustion process, BC can be
reduced by approaches other than direct reduction
using PM control devices. Process modification
and/or optimization can be an effective means of
reducing PM emissions. Some general examples of
process optimization include reducing the frequency
of mass transfer operations, improving operational
efficiency, and the proper use of dust collection
devices at the point of generation.
Cost values for these approaches are difficult to
estimate. Often, steps to improve operational
efficiency require only investments in
instrumentation or operator training to yield ongoing
reductions in fuel consumption and emissions of
all pollutants, including BC. Changes in fuel can be
more expensive and are incurred over the entire
period in which they are used, but are dependent
upon fluctuating and often highly localized market
conditions. However, for many situations, particularly
smaller boilers for which the costs of control
technology investment and operation would make
up a significant fraction of the system's initial capital
and operating cost, conversion to fuels (such as
natural gas) that generate lower PM and BC can be
less expensive than use of post-combustion control
equipment.
One specific example technique to reduce PM
emissions for existing boilers is a boiler tune-
up. Fuel usage can be reduced by improving the
combustion efficiency of the boiler. At best, boilers
may be 85% efficient and untuned boilers may
have combustion efficiencies of 60% or lower.
As combustion efficiency decreases, fuel usage
increases to maintain energy output resulting in
increased emissions. Lower combustion efficiency
also results in formation of PM constituents like BC
that are formed from incomplete combustion of the
fuel. The objective of good combustion is to release
all the energy in the fuel while minimizing losses
from combustion imperfections and excess air. A
tune-up can make a significant difference in energy
consumption and emissions levels.
9.6.2 Fuel Substitution and Source
Reduction Approaches for PM
The type of fuel and process has a great impact
on PM emissions from combustion. Coal, oil, and
natural gas are the most common fuels used. Of
these fuels, coal combustion generally results in the
highest PM emissions. As noted earlier, increased
use of biomass fuels may also lead to higher BC
emissions unless suitable control techniques are
applied.
Fuel substitution can be an effective means of
reducing PM emissions for many industrial fuel
combustion processes that generate process heat
or electricity. Switching to fuels that generate
lower levels of PM and BC per Btu can be a viable
alternative. In addition, fuel switching can lead to
cleaner and safer unit operation. However, there are
several factors to consider when evaluating whether
fuel switching would be an appropriate option.
Fuel-switching can often be implemented for a
lower capital investment than add-on control
technologies. In some situations, the age of a boiler
or space constraints may make fuel-switching
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more cost-effective. It is important to consider,
however, that the lower capital cost must be weighed
against a change in fuel prices, such as would be
incurred by switching from coal to natural gas. The
actual cost of converting to a different fuel that
reduces BC emissions must account for the cost of
installing the necessary fuel feed systems, fuel price
differential, and changes in non-fuel maintenance
and operational costs. Other considerations include
the cost of extraction and associated environmental
impacts. Even switching between types of coal can
lead to additional capital costs in the event new
coal pulverizers or a larger ESP is needed due to
differences in coal hardness and fly ash resistivity.
Fuel substitution for the purpose of reducing BC
emissions can also reduce emissions of SO2, NOX,
and CO, depending upon the characteristics of
the original and replacement fuels. A common
conversion is from coal to natural gas, with coal to
distillate oil an appropriate alternative. Switching
from distillate oil to natural gas is also a possible
approach for reducing BC emissions, but the
reductions in BC for such a change will be less than
when switching from coal to natural gas. Switching to
natural gas from either coal or oil will likely result in
significant SO2 emissions reductions.
When considering fuel switching to reduce BC
effects on climate, the impact on CO2 emissions is
obviously a consideration. Conversion from coal to
natural gas will reduce CO2 as well as BC. A further
alternative may be to switch to a biomass-based fuel
oil. The bulk of liquid biofuels appropriate for use
in boilers is in the form of biodiesel, although there
have been some evaluations of other biomass-based
fuel oils developed specifically for use in boilers
(Partanen and Allen, 2005; Adams et al., 2002).
9.7 Mitigation Approaches for
Stationary Sources Internationally
As discussed earlier in this section, stationary source
emissions of PM are generally considered to be
well-controlled in most developed countries due to
the operation of common control technologies such
as fabric filters and ESPs. The picture is different in
developing countries, where a number of specific
industrial source categories have been identified
as important contributors to BC emissions. The
source categories of concern vary by country and
region of the world. Mitigation opportunities exist in
these countries and regions because known control
technologies exist and have been demonstrated
to be effective. This section will address the source
categories that have been identified in the emissions
inventories as being major contributors to BC
emissions and for which known control technologies
exist. The source categories are brick kilns, coke
production/iron and steel production, power
generation and industrial boilers, and oil and gas
flaring.
9.7.1 Brick Kilns
Brick and masonry production in many developing
countries (such as China, India, Bangladesh, Vietnam,
Nepal, and Pakistan) has increased in recent decades
in response to growing urbanization and increasing
demand for construction materials. Currently, brick
production is estimated to be growing at a rate of
4% per year.9
Conventional brick kilns (such as bull's trench,
clamp, and intermittent downdraught kilns)
generally are operated by small-scale ventures in
rural areas, often with poor conditions for workers
(French, 2007; Gupta, 2003). Low-quality coal and
firewood are common fuels used in brick-making;
in some cases, even waste fuels such as used tires
are employed. These kiln designs have inefficient
combustion, leading to high emissions of both
greenhouse gases and PM (and associated local air
pollution health effects). The inefficient operation
of these kilns also leads to high fuel costs, and
kiln operations have been found to contribute to
localized deforestation when cheap firewood is
harvested in lieu of purchasing more expensive coal
to use as fuel.
The most basic BC mitigation technique is the
replacement of inefficient kilns with kilns having
improved, energy efficient designs, such as the
vertical shaft brick kiln (VSBK), the tunnel kiln, or
the hybrid Hoffman kiln (HHK). These kilns generally
require less than 50% of the fuel needed for a
conventional kiln (UNDP, 2007) and have been
estimated to reduce PM emissions proportionally.
Bond and Sun estimated that reducing emissions
by switching to a more efficient kiln design can be
cost-effective, in the range of $6.5 to $13 per ton
(2010$) of CO2-equivalent (based on 20-year GWP)
(Bond and Sun, 2005). China has taken steps over
the past decade to promote the transition to the
more efficient HHK in many areas. In Bangladesh,
the United Nations Development Program initiated a
$25 million project in 2010 to implement 15 energy-
efficient kiln demonstration projects over the next
five years (UNDP, 2010).
Under a "business as usual" scenario, global BC
emissions from brick kilns are expected to decline by
about 11% (428 to 381 Gg) over the 2005-2030 time
' See http://www.resourceeffidentbrKks.org/background.php.
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Chapter 9
period, reflecting a gradual introduction of more
efficient kilns. However, the technical mitigation
potential for this sector exceeds this projected
reduction. Given the rapid rate of urbanization
projected for coming decades in many countries, and
the high fuel cost and significant health and climate
impacts associated with uncontrolled brick kilns,
appropriate policy options, technical assistance, and
financial incentives could be considered to accelerate
the transition to more efficient brick kilns.
9.7.2 Coke Production/Iron and Steel
Production
Coke is a key input used in the production of iron
and steel. In the coking process, coal is heated to
very high temperatures for up to 36 hours in an
airless furnace, and volatile carbonaceous gases
are driven off. Modern plants minimize emissions
by capturing the coke oven gas and using it in a
separate chemical recovery process where it is
refined into by-products and usually burned for heat
production (RTI International, 2008). However, some
small-scale plants located in developing regions
or countries with economies in transition still do
not capture the carbonaceous emissions from coke
production. These plants in particular represent
potential BC mitigation opportunities.
The global demand for coke and steel has increased
significantly in the past two decades and is expected
to increase significantly in the future. Bond et al.
(2004) developed global emissions estimates for
BC from coke production of 380,000 Gg, based on
1996 data (about 8% of estimated global "contained"
emissions). It is acknowledged that this estimate
was highly uncertain due to the lack of information
regarding the number of polluting "beehive" or
"indigenous" plants currently in operation globally.
In the late 1990s, China was considered to have the
largest coke production capacity of any country
by far; and it continues to be responsible for more
than 60% of global coke production (based on
2008 data) and more than a third of global steel
production. Most coke production in China is
conducted by state-owned enterprises. However,
in 2004 it was estimated that smaller township and
village enterprises operating less capital intensive
"indigenous" plants were responsible for about 15%
of the coke production in China; it is assumed that
these smaller but uncontrolled plants are responsible
for a majority of BC emissions from the industry
(Dukan, 2010; Polenske and McMichael, 2002).
One mitigation option to reduce BC from coke
plants is simply to phase out smaller uncontrolled
operations. China has initiated policies to phase
out certain plants with uncontrolled emissions, but
the portion of the industry that has shut down or
consolidated, and the extent to which emissions
have declined to date, is not well characterized.
Other standard mitigation options may be
feasible depending on the size of the operation.
Emissions generated during the "pushing" of coke
from the oven to the quenching operation can
be captured by a moveable hood system and
then sent to a baghouse for control (with up to
98% PM reduction). Fugitive PM emissions can
be minimized during other stages of the coking
process with implementation of good combustion
practices, frequent maintenance, and proper work
practices. The use of a continuous opacity monitor
on the combustion stack can help identify ovens
that are in need of repair or maintenance (RTI
International, 2006). BC emissions associated with
larger coking operations could also be reduced via
implementation of an energy recycling program to
recover waste heat from the very high temperature
coking process. The recovered heat would be
converted to steam and used to power a generator
which in turn would help provide electricity needed
for plant operations, reducing the total amount of
coal needed run the plant and the associated BC
emissions (Polenske and McMichael, 2002).
Coke production and the iron and steel industry
are important contributors to BC emissions in
other regions as well. Bluestein et al. noted that
emissions from uncontrolled blast furnaces in the
former Soviet republics may have the potential to
contribute to BC levels in the Arctic (Bluestein et al.,
2008). However, additional information is needed
to improve global inventories of BC from coke
production. To the extent that existing sources in
China and other coke producing nations are not
recapturing exhaust gases, advanced technologies
are readily available to reduce emissions
significantly.
9.7.3 Power Generation and Industrial
Boilers
The worldwide demand for energy is projected to
increase substantially over the coming decades.
The 2011 International Energy Outlook projects
global energy demand to increase by 53% between
2008 and 2035. In countries outside the OECD, the
increase in demand over this period is estimated to
be 85%, while in OECD countries it is estimated to
be 18% (USEIA, 2011). With this increased demand
for power generation, increasing emissions of PM
and BC can be expected unless actions are taken to
implement energy conservation programs, to ensure
high combustion efficiencies in power plants and
industrial boilers, and to implement effective control
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technologies on new and existing facilities to the
maximum extent possible.
Internationally, many sources of power generation
(especially smaller power plants, industrial boilers,
and stationary diesel engines) continue to operate
without effective controls on PM and represent
current opportunities for mitigation. In 2001, it was
estimated that 20% of the power plants in China
were operating without effective PM controls
yet were responsible for 62% of the total PM2.5
emissions from power plants. In addition, many
industrial boilers in China are known to operate
only wet scrubbers and cyclones, which are effective
in capturing larger particles but have low fine-
particle removal efficiencies (Zhang et al., 2007). In
regions of the world where electricity from the grid
is unreliable or not available, there is a substantial
reliance on stationary diesel generators for power.
For example, it is estimated that diesel generators
in India account for as much as 17% of total power
generation (USAID, 2010a). Diesel generators are
also widely used in the Arctic region and contribute
to BC deposition locally (Quinn et al., 2008). Since
well-established control technologies are available
to control emissions from these power generation
sources effectively, additional emissions reductions
can be achieved. However, further investigation
is needed to determine the cost-effectiveness of
control options in specific locations.
9.7.4 Oil and Gas Flaring
Natural gas is a byproduct of the oil extraction
process and it is often treated as a waste gas and
disposed of rather than captured for economic use.
When not captured, it is either directly vented to
the atmosphere or it is burned through a process
called flaring. The combustion process during flaring
can be inefficient and characterized by a distinct
dark-colored, sooty plume. Oil and gas flaring and
venting leads to significant emissions of greenhouse
gases (especially methane) and a variety of other
air pollutants, including BC, hydrocarbons and toxic
air pollutants. Flaring can lead to significant health
impacts on nearby communities. BC emissions from
flaring are of particular concern if they can impact
areas of snow and ice in the Arctic region.
Global estimates of pollutant emissions from flaring
and venting are still quite uncertain. It has been
estimated that globally the natural gas wasted due
to flaring is about 5% of the total annual natural
gas consumption. In 2002, the World Bank started
the Global Gas Flaring Reduction initiative. Many
countries are now self-reporting flaring and venting
data. NOAA has also developed methodologies to
estimate flaring activity through the use of satellite
remote sensing data. Based on this information, the
countries with the highest estimated levels of flaring
are Nigeria, Russia, Iran, Iraq, and Angola (Buzcu-
Guven et al., 2010). More work is needed to improve
estimates of BC emissions from flaring. Mitigation
of venting and flaring activities will be on ongoing
challenge for the future. Reducing BC (and methane)
emissions from flaring and venting activities would
require expanded efforts to make use of the natural
gas for power generation on site or to capture the
gas so that it can be distributed and marketed.
There are clear economic incentives for this. EPA is
working with a number of governmental and private
partners to address these issues through the Global
Methane Initiative.
9.8 Technical and Research Needs
Emissions of BC from industrial sources, both
domestic and international, currently represent a
modest percentage of total BC emissions. In some
regions, such as Asia, industrial emissions are
more significant (almost a quarter of "contained"
emissions) (USAID, 2010a). It is expected that over
the next two decades, global emissions from the
industrial category will become a greater percentage
of "contained" global BC emissions as reductions
occur in other sectors. The reduction of BC and PM2.5
from industrial categories can be very cost-effective
when considering the substantial health benefits
they provide to local populations, in addition
to broader climate benefits. For these reasons,
controlling BC emissions from industrial sources
should remain a part of any overall BC reduction
strategy.
While this is the case, the emission factors and
emissions inventories for key sectors are recognized
by many experts to be uncertain, and there is a
need to improve PM2.5 and BC emission factors for
industrial sectors. In some cases, only a few source
tests may provide the basis for many emission
factors. It is difficult to measure BC emissions
that remain after control devices have treated the
exhaust emissions because it is difficult to measure
emission rates directly at the sources. BC emissions
are extremely difficult to measure under real-world
and field conditions. Our prior experience with
PM control clearly indicates that some ultra-fine
particles (BC and OC) are being captured in control
devices for larger particles, but reduction efficiencies
of control devices are generally considered to be
lower for sub-micrometer BC particles than for total
PM. To what extent is not well documented (Streets
et al., 2001).
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Chapter 9
To help develop improved PM and BC emission
factors, global inventories, and future year
projections, additional source test information needs
to be collected and evaluated for priority categories,
both in the United States and abroad. This research
would quantify the emissions of BC that pass
through existing control devices into the ambient
air, establish improved emission factors for different
source categories, and assess the engineering
modifications that can be made to these control
techniques to enhance their BC capture capability.
This could be facilitated by additional funding
for technical and research programs, and greater
collaboration between EPA, state governments,
industry groups, academic institutions, and
governments from other countries. Bilateral and
multilateral assistance programs can also play an
important role in evaluating the cost-effectiveness
of BC control measures in priority world regions.
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Chapter 10
Mitigation Approaches for
Residential Heating and Cooking
10.1 Summary of Key Messages
• In the developed world, residential combustion
is a small but potentially important source of
BC emissions. There are clear health benefits of
reducing residential wood smoke both indoors
and outdoors. The climate impacts depend on
the relative proportion of OC emissions, location
of emissions (over ice/snow) and the type of
wood-burning appliances used. Upgrading
old wood stoves in areas with snow and ice to
cleaner-burning appliances (particularly gas-
burning) appears to be the most effective strategy
to reduce BC and OC from residential wood
combustion (RWC).
- U.S. RWC is approximately 3% of the domestic
BC emissions inventory. Residential wood
smoke contains PM2.5, air toxic pollutants (e.g.,
benzene), CH4, CO2, OC, BC, and BrC.
- EPA is currently working to establish new or
revised new source performance standards
(NSPS) for all types of residential wood
heaters, including hydronic heaters, furnaces,
and wood stoves.
- Mitigation strategies for RWC sources
have generally focused on either replacing
inefficient units (wood stoves, hydronic
heaters) with newer, cleaner units through
voluntary or subsidized changeout programs,
or retrofitting existing units to enable use
of alternative fuels such as natural gas
(fireplaces). New EPA-certified wood stoves
have a cost-effectiveness of about $3,600/
ton PM2.5 reduced, while gas fireplace inserts
average $l,800/ton PM2.5 reduced (2010$).
- The Arctic Council Task Force on Short-Lived
Climate Forcers has identified wood stoves
and boilers as a key mitigation opportunity
for Arctic nations. The Task Force has
recommended countries consider measures
such as emissions standards, change-out
programs, and retrofits to reduce BC from
wood stoves, boilers, and fireplaces.
In the developing world, about 3 billion people
depend on rudimentary stoves or open fires
fueled by solid fuels (e.g., wood, dung, coal,
charcoal, crop residues) for residential cooking
and heating. This number is expected to increase
in the coming decades. Cleaner cooking solutions
have the potential to provide huge public health
benefits, and may be particularly important for
reducing regional climate impacts in sensitive
regions such as the Hindu Kush-Himalayan-
Tibetan region.
- According to the WHO, exposure to
cookstove emissions leads to an estimated 2
million deaths each year; indoor smoke from
solid fuels ranks as the six largest mortality
risk factor and the fifth largest disease
risk factor in poor developing countries.
Reductions in exposure to these emissions
likely represent the largest public health
opportunity among all the sectors considered
in this report.
- The BC climate impacts from cookstoves
are likely to be strong in a regional scope,
and additional source testing and modeling
is needed to clarify the composition of
emissions from these sources and their net
climate impact.
- Cookstove mitigation activity today is
difficult to quantify definitively: while the
EPA-led Partnership for Clean Indoor Air
(PCIA) reported that PCIA Partners sold
about 2.5 million stoves in 2010, it is likely
that 5-10 million "improved" stoves are sold
each year by commercial entities. In addition,
there are no reliable data on the quality or
performance of many of these stoves and
thus considerable uncertainty regarding the
benefits of their use. The full market of stoves
is on the order of 500-800 million homes (3
billion people); thus, significant expansion of
current clean cookstove programs would be
necessary to achieve large-scale climate and
health benefits.
Report to Congress on Black Carbon
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Chapter 10
Many improved cooking solutions exist, but
all face important supply, cost, performance,
usability, marketability and/or other barriers
to large-scale progress. The potential climate
and health benefits vary substantially by
technology and fuel.
o The performance hierarchy for improved
cooking solutions appears to be as follows,
in generally decreasing order for both costs
and emissions performance: 1) electricity;
2) clean fuels such as LPG or ethanol; 3)
advanced biomass stoves (e.g., forced air
fan or gasifier stoves); 4) rocket stoves;
and 5) other improved stoves. For all solid
fuel stoves (3 and 4), processing the fuel
into pellets or briquettes allows for greatly
improved combustion with significant
reductions to harmful emissions.
o Well-designed biogas may be the cleanest,
most climate-neutral (renewable) cooking
solution suitable for large-scale use; solar
stoves are ultimately the cleanest solution,
but have not yet demonstrated an ability to
reach large scales of sales or adoption.
A number of recent developments—including
the growth of a variety of promising
businesses and business models; innovations
in stove design, testing, and monitoring;
carbon financing; research quantifying the
health benefits of improved stoves; and new
country-based and global efforts to address
health risks—have created a real opportunity
to achieve clean cooking solutions at a global
scale.
o Over the past nine years, the EPA-led
PCIA has built a network of more than
540 Partners working in 117 countries to
increase the use of affordable, reliable,
clean, efficient, and safe home cooking
and heating technologies. PCIA Partners
sold approximately 2.5 million stoves in
2010, which may result in reduced indoor
air pollution exposures for more than 12
million people—primarily women and
children.
o Launched in September 2010, the
rapidly growing Global Alliance for Clean
Cookstoves (the Alliance) is led by the
United Nations Foundation and currently
has over 275 partners, including 28
countries and significant U.S. government
participation. The Alliance represents an
enormous opportunity to rapidly increase
the use of clean cooking solutions by
building on the past experiences and
successes of PCIA and other leaders in
this field (e.g., Shell Foundation, GIZ, SNV,
United Nations agencies, World Bank).
o The Alliance's interim goal is for 100
million homes to adopt clean cooking
solutions by 2020, with sales accelerating
through this period. Achieving this scale
of progress will not be easy—it will
require significant investments, demand
a coordinated global approach, and need
to be based primarily on sustainable
commercial businesses that produce high-
quality stoves and fuels that meet local
users' needs.
o Developing globally recognized
performance standards for stoves that
are widely accepted by the cookstove
community and adopted by country
governments could spur wider
development of clean cookstoves.
10.2 Introduction
Household energy use represents an extremely
important source of BC emissions worldwide,
accounting for 25% of the total global BC inventory.
In developed countries, most of these emissions
are associated with residential wood combustion
(RWC), generally for heating. Total emissions from
RWC in developed countries are estimated at about
4% of the total global inventory (311 Gg) and 16% of
total residential emissions worldwide. In developing
countries, emissions from residential combustion
are more often linked to widespread use of small
stoves for cooking and/or heating. These cookstoves
utilize a wider range of fuels, including coal, natural
gas, and dung as well as wood, charcoal, and
other biomass-related fuels. The emissions from
residential cookstoves represent a much larger
fraction of the global inventory, accounting for 21%
of total global BC emissions (1635 Gg) and 84% of
emissions from residential sources worldwide. The
variety of sources and fuels within the residential
category, and the significant differences between
developed and developing countries make this
sector among the most challenging from a
mitigation perspective. However, given the vast
number of people dependent upon residential
sources for everyday needs, such as heating and
cooking, this sector likely represents the biggest
opportunity for public health improvements through
reductions of BC and overall PM2.5.
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Mitigation Approaches for Residential Heating and Cooking
This chapter is divided into two parts. First, it
presents information regarding available mitigation
approaches for residential wood combustion in
the United States and other developed countries.
There are a number of cost-effective, advanced
mitigation technologies that are well known and
easily deployed; the biggest challenge remains
one of implementation and outreach. The chapter
then examines the technologies and approaches
available for reducing emissions from the residential
sector in developing countries, where the scale of
the problem is much broader, the range of sources
and fuels more complicated, and the challenges to
effective implementation much larger. It describes
the advanced cookstove technologies that are
currently available and their costs, and considers the
emissions reduction potential if these technologies
were adopted on a large scale.
10.3 Residential Wood Combustion in
Developed Countries
There are an estimated 29 million wood-burning
fireplaces, over 12 million wood stoves and hundreds
of thousands of hydronic heaters (also known as
outdoor wood boilers) throughout the United States.
Emissions from these appliances contain PM2.5,
toxic air pollutants, and other pollutants that can
adversely impact health and climate. The majority
of these emissions come from old, inefficient wood
stoves built before 1990. Wintertime wood smoke
emissions contribute to PM2.5 nonattainment and
localized problems in many areas in the United
States. For this reason alone, replacing inefficient
wood stoves and educating wood burners on proper
burn practices and stove operation are important
strategies for reaching domestic air quality goals. In
fact, there is far greater certainty about the public
health benefits of reducing residential wood smoke
emissions, both indoors and outdoors, than about
the net climate impacts, especially in light of the high
level of OC emissions from these sources.
10.3.1 Emissions from Residential Wood
Combustion
Incomplete combustion of wood results in emissions
of fine and ultrafine particles, including BC, BrC and
other non-light absorbing OC particles. Inorganic
materials, such as potassium, are also present
in lesser quantities as part of the mix of emitted
particles. In the United States, RWC contributes
over 350,000 tons of PM2.5 nationwide—mostly
during the winter months. Of this, approximately
21,000 tons is BC, which is about 3% of total U.S. BC
emissions. The key emitting source categories that
comprise RWC are wood stoves, manufactured and
masonry fireplaces, hydronic heaters, and indoor
furnaces. The 2005 PM2.5 inventory shows that cord
wood stoves contribute about 52%, fireplaces 16%,
hydronic heaters 16%, indoor furnaces 11% and
pellet stoves and chimineas (free-standing outdoor
fireplaces) the remaining 5%. Since 2005, the
popularity and use of outdoor hydronic heaters has
grown. As a result the emissions from these units
are growing and are of particular concern to many
areas, like the Northeast and Midwest.
In addition to PM25 and BC, wood smoke
contains toxic air pollutants such as benzene
and formaldehyde, as well as CH4, CO, and CO2.
Nationally, RWC accounts for 44% of polycyclic
organic matter (POM) emissions and 62% of the
7-polycyclic aromatic hydrocarbons (PAHs), which
are classified as probable human carcinogens.1
All of these pollutants are products of incomplete
combustion (PIC). These emissions are the direct
consequence of poor appliance design and
improper owner operation (e.g., using unseasoned
wood) leading to incomplete combustion of the fuel.
OC emissions from RWC generally far exceed the
BC emissions, making the OC/BC ratio relatively
large. However, different wood burning appliances
combust wood in varying ways, resulting in different
OC/BC ratios. In general, wood stoves have lower
OC/BC ratios than fireplaces (see Figure 10-1), and
also represent a significantly larger percentage of
the PM2.5 emissions inventory. The type of wood
burned also affects the amount of BC and OC
emissions.
Despite the relatively high OC/BC ratio from RWC in
general, it is important to consider the location of
these emissions. While OC emissions are generally
considered to have a cooling effect, OC emissions
over areas with snow/ice may be less reflective
than OC over dark surfaces, and may even have
a slight warming effect (see Flanner et al., 2007).
Significantly, the vast majority of residential wood
smoke emissions occur during the winter months;
the highest percentage of wood stove use is in the
upper Midwest (e.g., Michigan), the Northeast (e.g.,
Maine), and the mountainous areas of the Pacific
Northwest (e.g., Washington), where snow is present
a good portion of the winter months.
10.3.2 Approaches for Controlling
Emissions from RWC
Mitigation of RWC PM25 emissions generally involves
increasing the combustion efficiency of the source.
1 See EPA's 2005 National-Scale Air Toxics Assessment at http://
www.epa.gov/nata2005.
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Chapter 10
Wood burning appliances with
lower combustion efficiencies
tend to have higher emissions of
most pollutants than do those
with higher efficiencies. Due
to design, conventional wood
stoves, most fireplaces, and
outdoor hydronic heaters do not
burn wood efficiently or cleanly.
Mitigation strategies for RWC
sources have generally focused
on either replacing inefficient
units (such as wood stoves and
hydronic heaters) with newer,
cleaner units through voluntary
or subsidized changeout
programs, or retrofitting existing
units (such as fireplaces) to
enable use of alternative fuels
like natural gas. The United
States has been working to
establish emissions standards
for certain RWC sources, but it
takes time for such programs
to become effective, as they
depend on the turnover in
existing units. This is discussed
more fully below.
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To achieve the cleanest and
most efficient combustion, the
appliance needs to reach and
maintain a sufficiently high
temperature for all the necessary reactions to occur;
adequate time for those reactions; and enough
turbulence to ensure oxygen is available when and
where it is needed. EPA-certified wood stoves, wood
pellet stoves and Phase-2 qualified outdoor wood-
fired hydronic heaters2 are typically designed to
increase temperature in the firebox and to allow
for adequate outside air to mix long enough for
more complete combustion. The importance of the
combustion conditions within these home-heating
appliances, and the wood species used as fuel, in
determining the composition of the resulting wood
smoke is reflected by the observed variability in
measured OC/BC ratios discussed above.
In general, greater combustion efficiency leads
to reductions in the mass of direct PM emissions,
including BC, as well as reductions in emissions
of the gas-phase pollutants such as CO, CH4, and
the volatile PAHs. For example, in an EPA study
comparing a New Source Performance Standard
(NSPS)-certified wood stove to a traditional zero
2 For a list of such appliances, see http://www.epa.gov/burnwise/
owhhlist.html.
OC/BC Emission Ratios by Source Category and Fuel Type.
EPA)
clearance fireplace, the total PAH emission factor
was found to be up to twice as high for the fireplace
as for the more efficient stove burning the same oak
fuel (Hays et al., 2003). The same can be observed
for other pollutants depending on appliance type,
wood species, moisture content, and so forth.
A more efficient appliance also burns less wood
for the same heat output, leading to additional
emissions reductions. However, a recent wood stove
changeout study conducted by the University of
Montana showed significant reductions in emissions
of OC and levoglucosan (a wood-burning tracer)
but little or no change in BC emissions from the
changeout (Ward et al., 2011).
10.3.3 Emissions Standards for New Wood-
burning Units
EPA has authority to establish NSPS emissions
standards for new RWC sources, such as fireplaces,
wood stoves, and hydronic heaters. These standards
establish manufacturing requirements to limit
emissions from new units. Such standards can be
updated over time as new technologies become
available. Since 1988, EPA has regulated PMZ5
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Mitigation Approaches for Residential Heating and Cooking
emissions from new residential wood heaters
sold in the United States. The Residential Wood
Heaters NSPS (also referred to as the wood stove
NSPS) defines a wood heater as an enclosed, wood
burning appliance capable of and intended for
space heating or domestic water heating that meets
specific criteria, including an air-to-fuel ratio in the
combustion chamber averaging less than 35-to-l; a
usable firebox volume of less than 0.57 cubic meters
(20 cubic feet); a minimum burn rate of less than 5
kg/hr (11 Ib/hr) tested by at an accredited laboratory;
and a maximum weight of 800 kg (1,760 Ib). Many
types of sources are exempt from the existing NSPS,
including:
• Wood heaters used solely for research and
development purposes
• Wood heaters manufactured for export (partially
exempt)
• Coal-only heaters
• Open masonry fireplaces constructed on site
• Boilers
• Furnaces
• Cookstoves
The Residential Wood Heaters NSPS is unusual in
that it applies to mass-produced consumer items and
compliance for model lines can be certified "pre-sale"
by the manufacturers. A traditional NSPS approach
that imposes emissions standards and then requires
a unit-specific compliance demonstration would
have been very costly and inefficient. Therefore, the
NSPS was designed to allow manufacturers of wood
heaters to avoid having each unit tested by allowing,
as an alternative, a certification program that is used
to test representative wood heaters on a model
line basis. Once a model unit is certified, all of the
individual units within the model line are subject to
similar labeling and operational requirements.
EPA is currently in the process of revising the
Residential Wood Heaters NSPS. Specifically, the
Agency is considering tightening the air pollution
emission limits, adding limits for all pellet stoves,
reducing the exemptions, and adding regulations for
more source categories, including hydronic heaters
and furnaces. EPA expects to propose appropriate
revisions in 2012, and finalize revisions in 2013.
The tightening of the wood heater NSPS has the
potential to help reduce future residential wood
burning emissions throughout the United States.
10.3.4 Mitigation Opportunities for In-Use
RWC Sources
A fundamental limitation of the standards for
new sources discussed above is that they cannot
influence emissions from units that were purchased
prior to establishment of the NSPS. It can take a
long time for NSPS to actually reduce emissions,
depending on the rate of replacement of existing
units—and in many cases, these units can remain
in service for decades. Thus, alternative mitigation
strategies are needed to reduce emissions from
existing sources.
In 2004, a panel convened by the National
Academies of Science made several
recommendations to the EPA for improving air
quality management in the United States. One of
their recommendations was to develop and support
programs to address residential wood smoke. Since
2005, EPA has developed a residential wood smoke
reduction initiative that has various components
to support state, local, and tribal communities
in addressing their wood smoke challenges. This
initiative focuses on ensuring that wood burning
is as clean and efficient as possible to help reduce
emissions of harmful pollutants, the amount of fuel
used, and the risk of chimney fires from creosote
that builds up due to incomplete combustion. In
general, these programs were developed to reduce
PM2.5 and toxic air pollutants, but can be employed
to help reduce BC and other GHG (e.g., CH4 and
CO2) from RWC. The initiative has the following key
components.
10.3.4.1 Great American Wood Stove Changeout
Program
The hearth industry estimates that of the 12 million
wood stoves in U.S. homes today, 75% are wood
stoves built before 1990. EPA is working with
the hearth products industry and others to help
state, local, and tribal agencies create campaigns
to promote replacement of old wood stoves and
wood-burning fireplaces with new, cleaner-burning
and more energy efficient appliances. Programs vary
from one community to another, with some areas
focusing on changing out old wood stoves and
others on retrofitting open fireplaces with cleaner
burning options (e.g., gas stoves). The campaigns
are typically led by local government or non-profit
organizations at the county or regional level.
Residents of participating communities generally
receive incentives such as cash rebates, low/no
interest loans and discounts to replace their old,
conventional wood stoves and fireplace inserts
with cleaner-burning, more efficient EPA-certified
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Chapter 10
New Wood Stoves and
Pollution Reduction
EPA estimates that every 1,000 old wood stoves
changed out to cleaner burning hearth appliances will
result in annual pollution reductions of
• 815 tons of C02
• 53tonsofCH4
• 27 tons of PM2.5
• 4 tons of toxic pollutants
• 14 tons of OC
• 1.6 tons of BC
(Numbers generated using EPA's Wood Stove and
Fireplace emissions calculator: http://www.epa.gov/
burnwise/resources.html and EPA's speciation profile data
base.)
gas, pellet, electric, wood stoves and fireplaces or
even geothermal heat pumps. A new EPA-certified
wood stove, new flue, and professional installation
cost, on average, $3,500
(2010$). Some areas have
provided cash incentives to
low-income participants only,
while others have provided
incentives to everyone in the
community. The local agency
leading the replacement
program will sometimes include
weatherization programs which
insulate homes to help reduce
heat loss and reduce fuel
consumption. Households that
participate in these programs
are required to surrender their
old wood stoves to be recycled.
$12,000 —|
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$4,000 -
Some of the benefits of
replacing inefficient wood
stoves include:
• Reduction in PM2.5 and
toxic air pollutants (e.g.,
benzeno(a)pyrene) by 70%
• Reduction in indoor PM2.5
emissions by 70% according
to University of Montana3
• Improvement in energy efficiency by 50%, using
one-third less wood
• Reduction in CH4, BC, and CO2 from improved
combustion efficiency and use of less fuel wood.
A variety of examples of state and local efforts to
reduce emissions from older appliances are available
at EPA's Burn Wise website: http://www.epa.gov/
burnwise/casestudies.html.
EPA's wood stove changeout effort has focused
primarily on counties at or near nonattainment
for PM2.5 where wood smoke is an important local
source. EPA estimates that through 2011, the Great
American Woodstove Changeout Program has
helped changeout or retrofit nearly 24,000 wood
stoves and fireplaces in 50 areas. From 2010 on, this
program is anticipated to reduce approximately
370 tons of PM25 and 63 tons of hazardous air
pollutants (HAPs) each year after 2010, providing
approximately $120 million to $330 million (2010$) in
estimated health benefits in the U.S.4
The best available cost-effectiveness information on
residential wood smoke mitigation comes from a
EPA-certified wood stove
Wood pellet stove
Gas stove
Figure 10-2. Cost Per Ton PM2.5 Reduced for Replacing Non-EPA-
Certified Wood Stove with EPA-Certified Woodstove (in 2010$).
(Source: U.S. EPA, based on data from http://www.mamma.org/visibility/
ResWoodCombustion/RWC_FinalReport_121906.pdf)
3 For more information, see: http://www.ncbi.nlm.nih.gov/
pubmed/18665872.
4 These estimates reflect national average benefit-per-ton estimates
for directly emitted carbonaceous particles from area sources from
http://www.epa.gov/air/benmap/bpt.html (data accessed February
2011) and Fann, Fulcher, Hubbell 2009 methodology. These
estimates have been inflated from 2006$ to 2010$.
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Chapter 10
Education and Outreach: EPA's Burn Wise Campaign
Perhaps one of the biggest opportunities to reduce wood smoke emissions, including BC, lies in the hands of those who
burn wood, regardless of the type of appliance they own. How wood stoves are operated and what is burned are as
important as the type of stove used. EPA has heard from state, local, and tribal governments and from the public that even
people who own an EPA-certified wood stove are often times burning "green" unseasoned wood, trash, and/or improperly
operating their appliance, resulting in high wood smoke emissions.
In October 2009, EPA launched an education campaign called Burn Wise to promote responsible wood-burning and
to educate users on the connection between what they burn, how they burn, and the impacts on their health and
the environment.The campaign provides a website (www.epa.gov/burnwise), fact sheets, posters, and public service
announcements. EPA has coordinated with the hearth products industry, chimney sweeps (Chimney Safety Institute of
America), and other partners on the development and implementation of the campaign.
Getting people to change their habits and behaviors, including their wood burning practices, is typically not a trivial or
inexpensive task. Equally challenging is measuring the effectiveness of social marketing or education campaigns like Burn
Wise. However, EPA does believe the benefits, particularly the public health benefits, are worth it, and that some methods
are more effective than others. For example, Environment Canada implemented a"Burn It Smart"campaign that included
conducting community based workshops.The workshops were targeted in areas where government officials believed
heating with wood was very common. Even though they did not calculate emissions reductions, a follow-up survey of
174 people indicated that
• 3% percent of the respondents said the workshops brought about positive change on how they burned wood
• 34% have updated their wood burning appliances; 90% of those chose EPA approved appliances
• 41% of those surveyed have changed out or intend to change out their old wood burning appliances for cleaner
technology.
Wood Burning Curtailment Programs: One of
the most effective ways a community can reduce
wood smoke is by developing a mandatory
curtailment program or instituting "burn bans."
Some communities implement both a voluntary and
mandatory curtailment program depending on the
severity of their problem. Curtailment programs
often have two stages, with Stage 1 allowing EPA-
certified wood stoves to operate and Stage 2
banning all wood burning appliances, unless it is
the homeowner's only source of heat. Although
curtailment programs are not always popular with
the public, this measure can be highly effective
at reducing wood smoke. As an example, the
Sacramento Air Quality Management District's Stage
2 program, implemented in 2008-2009, reduced
PM2.5 levels by 12 ug/m3. The cost effectiveness was
estimated to be approximately $6,300 - $11,100
per ton of PMZ5 reduced (2010$) (SMAQMD, 2009).
To increase the likelihood of success, curtailment
programs should include a forecasting and public
notification system. In addition, an enforcement
component is important to ensure the public takes
the program seriously.
Removal of Old Wood Stove Upon Re-Sale of a
Home: Old wood stoves are usually made of metal,
weigh 250 to 500 pounds, last for decades, and
can continue to pollute for just as long. As a result,
homeowners are less likely to replace old stoves
with a new, cleaner burning technology or remove
the old stove, especially if they are not using it. To
help get these old stoves "off-line," the state of
Oregon and some local communities in other states
have required the removal and destruction of old
wood stoves upon the resale of a home. Specifically,
this requirement has proven very effective in
locations like Mammoth Lakes, CA and Washoe
County, NV.7
10.3.6 Wood Smoke Reduction Resource
Guide
In October 2009, EPA released a resource guide
called Strategies for Reducing Residential Wood
Smoke8 that was written for state, local, and tribal
air pollution control officials so they would have
a comprehensive list of strategies to help reduce
wood smoke from residential heating. The guide
provides education and outreach tools, information
on regulatory approaches (e.g., burn bans) to
reduce wood smoke, as well as voluntary programs
to change out old, inefficient wood stoves and
7 For more information, see: http://www.gbuapcd.org/
rulesandregulations/PDF/Reg4.pdf.
8 http://www.epa.gov/ttncaaal/tl/memoranda/strategies-
doc-8-ll-09.pdf.
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Mitigation Approaches for Residential Heating and Cooking
fireplaces. It also notes the upcoming wood heater
NSPS has the potential to help reduce future
residential wood burning emission throughout the
United States. Several state and local communities
have effectively implemented residential wood
smoke control strategies and have significantly
reduced harmful wood smoke pollution. For example,
Lincoln County, MT and Sacramento Metropolitan
Air Quality Management District have encouraged
comprehensive wood smoke reduction strategies
to help these areas clean the air and protect public
health.
10.4 Residential Cookstoves in
Developing Countries
More than 3 billion people worldwide cook their
food or heat their homes by burning biomass
(e.g., wood, dung, crop residues, and charcoal) or
coal in polluting and inefficient traditional stoves
(International Energy Agency, 2010). As discussed
in Chapter 4, BC emissions from these sources are
estimated to account for 21% of the total global
inventory. This use of solid fuels also represents
a significant part of energy use in developing
regions—including nearly 50% of total primary
energy supply in Africa, and about 27% in India
(International Energy Agency, 2009). Use of biomass
and waste in developing nations—nearly all of which
is for household cooking and heating—accounts
for about 60% of global renewable energy use
(International Energy Agency, 2009, Annex A). About
82% of those who rely on traditional biomass fuels
for cooking live in rural areas; however, in Sub-
Saharan Africa, nearly 60% of people living in urban
areas also rely on biomass (International Energy
Agency, 2010).
As discussed in Chapter 3, several decades of
research document the significant risks to public
health associated with traditional cookstoves.
Exposure to cookstove emissions leads to an
estimated 2 million deaths each year and ranks as the
sixth highest mortality risk factor and fifth highest
disease risk factor overall (in terms of disability
adjusted life years—DALYs) in poor developing
countries (World Health Organization, 2009).
Emissions from cookstove use have been linked to
adverse respiratory, cardiovascular, and neonatal
outcomes and to cancer (Smith et al., 2004). Of the
total global mortality associated with exposure to
cookstove smoke, Sub-Saharan Africa, China and
India each account for approximately 25-30%, with
the remainder of deaths attributable to cookstoves
occurring elsewhere in Asia and Latin America.
The contribution of this source category to
emissions of BC and other aerosols has been the
focus of growing interest, especially in terms of
impacts on sensitive regions such as the Himalayas.
A recent study on BC emissions from cookstoves
in northern India indicated that cooking with solid
fuels is a major source of ambient BC in the region,
with peak ambient BC concentrations of about 100
ug/m3 (Rehman et al., 2011). Furthermore, the study
indicated that OC concentrations (which exceeded
BC concentrations by a factor of five) contained
significant absorbing BrC, and suggested that
previous estimates of atmospheric solar heating
in the region due to particles from cookstove
emissions should be increased by a factor of two or
more. However, there remains significant uncertainty
about the extent of BC and associated emissions
from cookstoves, and the effect of those emissions
on climate. Given the complex emissions mixture
resulting from cookstove use, further study is
needed to pinpoint the most beneficial strategies
for reducing BC emissions from this source.
Unquestionably, however, this sector represents the
area of largest potential public health benefit of any
of the sectors considered in this report. Mitigation
of emissions from cookstoves offers a tremendous
opportunity to protect health, improve livelihoods,
and promote economic development—particularly
for women and children. For this reason alone,
irrespective of the additional climate benefits that
may potentially be achieved, mitigation of cookstove
emissions is a pressing priority.
Mitigating BC emissions from cookstoves depends
first on identifying technologies that are proven
effective in reducing BC emissions, and second
on encouraging adoption of these technologies
on a large scale. As discussed below, very few
improved stoves have been designed specifically
to reduce BC emissions, and while some improved
technologies are emerging, no advanced stoves
that burn solid fuel have yet been adopted on
a broad scale (though LPG has been widely
disseminated as a clean cooking fuel, and China
(see below) implemented a very large earlier stoves
program using intermediate stoves). The problem
is complicated by the fact that both the impacts
of cookstoves and the solutions are regionally
dependent. Specifically, the extent of achievable
BC reductions, and the impact of those reductions,
will depend on the type of stove, the type of fuel
used, and the location of emissions. Improved
cookstoves and fuels must satisfy the needs of local
users, enabling them to cook local foods at the
time and in the manner they prefer, using the fuels
that are available and affordable. Given the array of
different technologies and fuels currently in use, and
the sheer number of sources involved, mitigation
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Chapter 10
of BC emissions from cookstoves represents an
enormous challenge. However, given their significant
contribution to the global inventory of emissions,
and the increasing availability of cost-effective
and locally appropriate solutions and several other
factors noted in section 10.4.3, this sector represents
one of the most promising opportunities for
mitigation of BC internationally.
10.4.1 Emissions from Cookstoves
Currently, residential cookstoves contribute
approximately 21% of the global BC emissions
inventory, with emissions concentrated in Sub-
Saharan Africa, China, India, and other developing
regions of Asia. Dependence on traditional biomass
fuels is highly correlated with poverty; countries with
higher household income also tend to have a higher
share of modern fuels for residential consumption.
While the percentage of people relying on traditional
biomass fuels for basic household energy needs is
expected to decrease in most areas over the coming
decades, the aggregate number of people relying
on biomass for cooking and heating is expected
to increase by 100 million people by 2030 due to
population growth (International Energy Agency,
2010). IEA projects that the fastest shift toward
modern fuels will occur in India, and the slowest shift
will occur in Sub-Saharan Africa (International Energy
Agency, 2010). The impact of these changes on
emissions is still unclear: as discussed in Chapter 7,
under most scenarios, residential emissions are
projected to decline significantly by 2030 and further
still by 2050 (Streets et al., 2004). However, a decline
in emissions, and the rate of decline will depend
on rates of adoption of cleaner fuels and cooking
technologies described below, and some regions
may experience near-term increases in emissions.
10.4.2 Technologies and Approaches for
Controlling Emissions from Cookstoves
Because cooking is such a variable, individual-
specific activity, there are many complexities
related to achieving reductions in BC emissions
from improved cookstoves. The type of fuel and its
moisture content, the type of stove, the purpose
for which it is used (heating vs. cooking), and the
manner in which the stove is tended all affect
composition of emissions (MacCarty et al., 2008).
Cooking practices vary both daily and seasonally due
to variation in available foods and fuels, and variation
in fuel quality. Additionally, there may be significant
variation in the efficiency and durability of stoves,
even those that are mass produced.
There is currently no formal definition of what
constitutes an "improved" stove. In the past,
"improved" stoves typically meant low-cost, locally
made stoves aimed at improving efficiency and
reducing fuel use. A primary motivation for the
use of improved stoves was to reduce demand for
fuel wood, thereby reducing pressures on forests
as well as the time spent by women and children
gathering fuel (Graham et al., 2005; Partnership for
Clean Indoor Air, 2005). However, not all such stoves
functioned as intended. For example, stoves that
have a large amount of heated mass, such as the
Lorena stove, may remove smoke with a chimney,
prevent burns, and help warm a house, but may not
save fuel compared with an open fire (USAID, 2007).
Over the last ten years, a new suite of more effective
stoves has been introduced to the marketplace. As
a group, these new improved stoves are designed
to be much more efficient and clean (as well as
safe), and utilize a variety of different technologies
and fuels. Most are produced locally for the
nearby market, while there are a few that are mass
produced internationally and can be shipped
anywhere in the world. The stoves span a wide
range of cost, durability, and performance, and
are designed for different types of staple foods.
Importantly, however, these stoves are generally
designed to reduce fuel use and emissions of
PM2.5 and CO (as proxies for the broader suite of
emissions from these stoves). Few of the stoves
currently on the market were designed to reduce
BC specifically (the new Turbococina stove is an
exception). In laboratory settings, most of these
stoves achieve PM2.5 reductions of 40% to 70%.
Results from MacCarty et al. (2008) indicate that
some non-advanced stoves may not substantially
reduce BC emissions, but some gasifier and forced
draft (or "fan") stoves significantly reduce BC
emissions compared to the open fire. Field testing
has begun to determine whether stoves perform
as well in actual (real-world) conditions as in the
laboratory. Results from field tests indicate that
stove performance under actual conditions varies
(Roden et al., 2009; USAID, 2011). Much more such
testing is needed, as well as additional research and
development in stove design to determine if the
stoves are reducing BC in addition to total PM2.5.
Among the new technologies now emerging on
the cookstove market, there are a few advanced
solutions that have been shown to reduce PM2.5
by 90% or more in laboratory settings; the limited
lab testing performed on these stoves to date
indicates that some stoves reduce BC by a similar
percentage (MacCarty et al., 2008). These new
technologies include advanced forced-draft stoves
and "gasifier" stoves (Roth, 2011) that use various
solid biomass fuels (including wood, pellets, crop
residues, etc.); biogas stoves; and liquid-fuel stoves
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Figure 10-4. TheTurbococina Stove, which burns wood,
is made of a stainless steel cylinder fitted with 10 air
injectors, an internal fan that runs on electricity, and a steel
plate that regulates the air flow to improve combustion
efficiency. This stove is currently available primarily for
institutional settings. (Photo courtesy of Rene Nunez
Suarez, Turbococina)
that burn ethanol, plant oil, or other biomass fuels.
It is important to note, however, that few of these
stoves are widely commercially available and their
performance in the field has not been fully evaluated.
Some stoves require electricity to drive a fan (see
Figure 10-4), while others have been designed to
convert waste heat to electricity to power a fan,
which enables excellent emissions performance
(including BC emissions reductions) without the
need for access to electricity (see Figure 10-5).9
Some of these stoves are being further tested for
emissions in both the lab and the field. These new
stove technologies have the potential to reduce
emissions from cookstoves nearly to the levels of
clean fuels such as LPG (Wilkinson et al., 2009), but
many (though not all) require specific and/or highly
processed fuels, which increases the total cost of
use (Venkataraman et al., 2010). In some cases, clean
fuels (ethanol, biogas) are not widely commercially
available for cooking. Forced-draft stoves that do
not generate their own electricity require access to
another electricity source, which can be costly. Even
battery-powered fan stoves require intermittent
access to electricity. Some gasifier stove designs use
natural draft (natural convection) and do not require
a fan.
While the basic outlines of lab and field tests have
been in place for decades, it is only in the past five
years that organizations funding household energy
interventions have begun requiring emissions
pre-testing, or that performance benchmarks
(even informal ones) have been established.
Recent testing in both lab and field settings (see
below) demonstrate that this new generation of
stoves is achieving real and measurable results.
Developing globally recognized standards that are
widely accepted by the cookstove community and
adopted by country governments could spur wider
development of clean cookstoves.
Based on available performance and cost data,
currently available technologies exhibit a wide range
of performance. These options include:
• electric stoves
• gas and liquid fuels
• processed solid fuels
• advanced biomass stoves
• rocket stoves
9 These stoves may also soon be able to reliably generate enough
electricity to be used for other purposes (e.g., lights or cell phones),
which could increase consumer demand. The change in emissions
with the new stove would depend in part on the extent to which
overall stove usage increased due to demand for these extra
services (Venkataraman et al., 2010).
Figure 10-5. Woman Prepares Banku on a BioLite
HomeStove in Kintampo, Ghana. BioLite stoves convert
their waste heat into electricity, which is used to power
the fan and can be used to charge a battery for a mobile
phone (as shown in photo), LED light, or other low-power
purpose. (Photo courtesy of Jonathan Cedar, Biolite)
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• simple stoves
• solar cookers
• behavioral and structural solutions
10.4.2.1 Electric Stoves
Cooking with electricity produces zero emissions
within a household, and therefore is highly effective
at reducing personal exposures of stove users.
However, from a broader public health or climate
perspective, emissions associated with the increase
in power production must also be considered. The
ongoing electricity costs for these stoves can vary
substantially by region.
10.4.2.2 Gas and Liquid Fuels
Switching from solid fuels to gaseous or liquid fuels
is often the easiest means of dramatically lowering
emissions from cooking. In laboratory testing, the
Aprovecho Research Center (Aprovecho) found
that using liquid petroleum gas (LPG) decreased
the amount of energy used by 69%, the mass of
Figure 10-6. CleanCookEthanol Stove. Currently
these stoves are being deployed in refugee camps in
Ethiopia and Kenya, with over 3,800 stoves already in use.
Deployment of additional stoves is dependent partly on
the availability of ethanol. (Photo courtesy of Harry Stokes,
Project Gaia)
fuel used by 89%, particle emissions by over 99%,
CO emissions by 98%, and time to boil by 40%, as
compared to cooking over an open fire (MacCarty
et al., 2010). Field research in Guatemala showed
that LPG stoves could reduce indoor 24-hr PM2.5
concentrations by over 90% (Naeher et al., 2000).
Liquid fuels such as ethanol, kerosene, and plant oils
are also options (see Figure 10-6). Aprovecho's lab
tests found that cooking with well-made ethanol or
kerosene stoves decreased the mass of fuel used by
75% and 82% respectively and particle emissions by
over 99% for each; CO emissions by 92% and 87%
(MacCarty et al., 2010).
Biogas derived from waste biomass is potentially as
clean as LPG, and in addition it is renewably derived
(reducing CO2 impacts) and requires no distribution
infrastructure. Emissions testing of biogas stoves to
date suggests that these stoves perform significantly
better than solid fuel/stove combinations with
regard to emissions of methane, CO, VOC, and CO2
(Smith et al., 2000b). Plant oils are another liquid
fuel being used today for cooking, but independent
testing results for stoves using these fuels have not
yet been published. Stoves using gas and liquid
fuels involve an upfront cost of $5 to $50 per stove,
as well as an ongoing cost for the fuel that varies
substantially by region, fuel, and changing economic
conditions. LPG stoves can also require significant
deposits on the cylinders, another serious barrier
for the very poor. It is also important to note that
poorly made kerosene stoves in particular pose
safety concerns, including the potential for severe
burns and injury associated with accidental fires
(Peck et al., 2008).
10.4.2.3 Processed Solid Fuels
For much of the developing world, the advanced
solutions described above may be unavailable or
simply too costly to use in the near-term. Stoves
utilizing processed solid fuels in the form of
charcoal, pellets, prepared wood, and briquettes,
may be more accessible, and these can also
represent very clean solutions. However, like the
clean fuels noted above, using processed fuels also
involves an ongoing operating cost, which may
serve as a barrier for these solutions, especially
in regions where fuel wood can be collected
free of charge. However, in markets where fuel
is purchased, stoves that increase combustion
efficiency by 50% are often the easiest stoves to
market, since the consumer can expect a quick
payback period on the initial investment.
Charcoal is the most common processed solid fuel
used today. A number of charcoal stove models
are available (see Figure 10-7). Lab tests of charcoal
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(a)
Figure 10-7. Charcoal Stoves, (a) Charcoal Stove by Burn Design. (Photo courtesy of Peter Scott, Burn
Design) (b) Charcoal Zoom Versa Stove produced by EcoZoom. (Photo courtesy of Ben West, EcoZoom) (c)
Prakti Charcoal Stove. (Photo courtesy of Mouhsine Serrar, Prakti Design) (d) Envirofit CH-4400 Charcoal
Stove. (Photo courtesy of Envirofit) (e) Toyola Charcoal Stove in use in a village in Ghana. (Photo courtesy of
Suraj Wahab, Toyola Energy)
stoves for climate forcing emissions found that these
stoves—relative to an open fire—reduced the BC/
OC ratio somewhat, and reduced total particles
by about two thirds (MacCarty et al., 2008). It is
important to note that the laboratory emissions
tests do not account for emissions in the charcoal
production process, which is highly inefficient and
polluting, with significant net climate impacts (Bailis
et al., 2005). Aprovecho has tested many charcoal
stoves for PM2.5, CO, and fuel use, finding that PM2.5
emissions were 90% lower than for a 3-stone fire
and fuel use savings ranged from 45% to 65%.
Most charcoal stoves cut time to boil, though only
modestly. However, CO emissions increased for all
stoves except one (MacCarty et al., 2010). In 2007,
EPA tested two charcoal stoves and found that
relative to a 3-stone fire, PM2.5 emissions from the
charcoal stoves fell by over 90% from a hot start
but increased when operated from a raw, cold start,
and both stoves increased CO emissions (Jetter and
Kariher, 2008).
Creating pellets from biomass or briquettes from
either coal or biomass can lead to substantial
improvements in efficiency and emissions when
pellets are burned in well-designed stoves. The
Oorja stove (developed by BP and now owned by
First Energy of India) is an example of a very clean-
burning pellet stove—in this case the pellets are
made from crop residues by a partner company.
More than 400,000 Oorja stoves have been sold
and between 250,000 and 350,000 are in use every
day. However, given the cost of pellets, this stove
competes with LPG. Other examples include project-
based work that have developed briquettes from
waste biomass (Haiti, Ghana, and Uganda), stoves
designed to burn pellets made from locally available
waste biomass (West Africa and elsewhere), and a
stove that burns rice hulls (Philippines), though EPA
is not aware of any examples where this work has
been carried to a large scale. With regard to coal
cooking, laboratory measurements indicate that
the combination of using improved stoves with
processed coal briquettes could have a dramatic
impact on aerosol emissions. Zhi et al. measured
reductions in particles of 63%—with OC decreasing
61% and BC decreasing 98%. This reduced the BC/
OC ratio by about 97%, from 0.49 to 0.016 (Zhi et al.,
2009).
10.4.2.4 Advanced Biomass Stoves
There are two types of advanced biomass stoves
that can achieve high levels of performance: forced
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Figure 10-8. Philips Woodstove (forced draft)
Manufactured in Lesotho. (Photo courtesy of Stephen
Walker, African Clean Energy Ltd.)
draft and gasifier stoves. Gasifier stoves can be
forced-draft or natural draft. These stoves can burn
processed or raw biomass, though it is likely the case
(field testing data forthcoming) that those using
processed fuels will perform better in the field, since
processed fuels eliminate a major variable in real-
world use of the stoves. It is also likely that lab and
field emissions test results will be more consistent
for stoves that burn processed fuels. Lab testing of
advanced biomass stoves to date generally confirms
that these advanced biomass stoves can achieve
remarkable emissions reductions—up to 93%
lower than traditional stoves (Venkataraman et al.,
2010). One study found that these stoves achieved
substantial reductions in both overall particles
and BC specifically, with the fan stove significantly
reducing particle emissions and the gasifier stoves
reducing total particles by about two-thirds (as well
as reducing the BCOC ratio). The study also showed
that the fan stove was able to reduce time to boil, at
least under the lab conditions (MacCarty et al., 2008).
Under Aprovecho's broader lab testing, forced draft
fan stoves all reduced (relative to a 3-stone fire) fuel
use (by 37% to 63%), CO emissions (in all cases by
over 85%), PMZ5 emissions (from 82-98+%), and time
to boil (11% to 65%). Similarly, the gasifier stoves
tested by Aprovecho saved on fuel use, reduced
CO emissions, achieved dramatic reductions in
particle emissions (with one exception), and cut
the time to boil, though generally all to a lesser
extent than the fan stoves (MacCarty et al., 2010).
In EPA's 2007 testing, the one advanced fan stove
tested (Philips) had the best overall performance
and the lowest pollutant emissions, reducing
emissions of key pollutants such as PM2.5 and CO
by about 90%. Notably, of the wood burning stoves
tested, this stove was also the one that required
the least attention to operate (Jetter and Kariher,
2008), although it required fuel with short (<10 cm)
lengths (see Figure 10-8). A forthcoming study from
investigators in India of recently completed field
testing of two forced draft cookstoves and indicated
that both stoves substantially reduced BC emissions
in both the breathing zone (85% and 49% BC
reduction respectively, compared to the traditional
mud cookstove) and in the plume zone (86% and
64% respectively) (Kar et al., 2012). Indoor cooking-
time BC concentrations were reduced to 5-100 u.g/
m3 by the top-performing forced draft stove (as
compared to 50-1000 u.g/m3 for the mud stove).
It is important to note that, while very promising
in terms of performance, most of these models are
still in the research and development stage, though
a few have been introduced in the market today.
These stoves are typically more costly than other
biomass stoves, currently costing in the range of
$25-100 per unit (plus any processed fuel costs,
which can be substantial10), though prototypes
for newer versions have been developed that
manufacturers estimate will cost in the $40-60 range
at full production.
10.4.2.5 Rocket Stoves
Where advanced stoves are not widely available
in the marketplace, or are not affordable, rocket
type stoves are typically the most efficient and
clean biomass-burning alternative (see Figure 10-9).
Rocket stoves have a combustion chamber designed
to allow for better mixing of combustion gases and
higher combustion temperatures which slightly
improves combustion efficiency (compared to the
open fire) and reduces emissions. Additionally,
rocket stoves have substantially better heat
transfer efficiency, so as they save fuel, they reduce
emissions (for a given cooking task). Thus, rocket
stoves substantially reduce emissions without relying
on electricity or other sophisticated components.
However, rocket stoves have not been designed to
date to reduce BC emissions. MacCarty et al. (2008)
found that the rocket stove reduced total particle
emissions by about 40%, but that nearly all of the
emissions reductions were of organic matter; BC
emissions for this stove did not decrease (and thus
the BCOC fraction increased dramatically). The
study also showed that the rocket stove was able to
10 For example, the pellets for the Oorja stove in India cost roughly
7 Rupees (~15ct) per kilogram of pellets.
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(b)
Figure 10-9. Rocket Stoves, (a) EnvirofitGSSOO Biomass Cookstove in Use in Tanzania. (Photo courtesy
of Nancy Ryden, Envirofit) (b) Zoom Dura Biomass Cookstove produced by EcoZoom. (Photo courtesy of
EcoZoom)
reduce time to boil, though to a lesser extent than an
advanced fan stove.
Aprovecho tested a wide variety of rocket stoves and
found important variability in performance indicating
that design is critical. Most saved significantly (but
not equally) on fuel use (26% to 51% savings relative
to a 3-stone fire), though two failed to cut fuel use.
All rocket stoves cut CO emissions by 70% or more,
while performance on PM2.5 emissions varied much
more widely (one actually increased emissions), with
60% of those tested achieving reductions of over
50%. Some of the rocket stoves actually increased
the time to boil, though most cut it modestly
(MacCarty et al., 2010). In EPAs 2007 testing, several
non-advanced wood stoves were tested and results
varied depending on the design and the stage of
operation. Generally, emissions were lower than the
3-stone fire, with faster times to boil. For example,
the UCODEA wood stove—now called Ugastove—
reduced PM2.5 and CO emissions by 48% to 65%
when operated at high power, and 35% to 50% at
low power (Jetter and Kariher, 2008).
The U.S. Agency for International Development
(USAID) recently conducted extensive field testing
of five non-advanced biomass stoves in the Dadaab
refugee settlements in Kenya. They tested for fuel
use, time to boil, and several user preferences—but
did not test for BC or any other emissions—and
concluded that "all five tested stoves outperformed
the open fire, requiring significantly less fuel to
cook the test meal....with savings ranging from
32% to 65" (USAID, 2010b). Additional testing of
two manufactured rocket stoves by Columbia
University researchers demonstrated "substantial and
statistically significant fuel savings relative to the
three-stone fire" (38% and 46% on average, for the
two stoves), but further stressed that fuel savings is
just one factor that affects suitability of any given
stove in a particular community. Other relevant
factors include stove size, ease of use, and cooking
time (Adkinsetal., 2010).
USAID (2011) also completed a recent round of
cookstove testing in the field in Uganda that did
examine BC and other emissions and factors,
comparing a traditional stove to a leading rocket
stove. The study found that with the greater fuel
efficiency of the rocket stove (42% savings were
measured) came lesser emissions of PM40 and
carbon monoxide. However, the rocket stove (which
was not designed to reduce BC emissions) had
more than twice the fractional BC content in its PM
emissions (15.5%) compared to the traditional stove
Recent field testing in India (Kar et al., 2012) noted
above found preliminary results showing that the
natural draft stoves had much wider variation in
performance than the forced draft stoves, and did
not achieve nearly as great reductions in BC. For
example, these natural draft stoves reduced BC
emissions in the breathing zone by a factor of 1.5
on average (22% to 55%), as compared to a factor
of 4 for forced draft stoves. However, BC reductions
varied significantly among natural draft stoves:
only micro-gasification stoves were shown by Kar
et al. (2012) to be effective in reducing BC, while
other models occasionally emitted more BC than a
traditional cookstove. BC emissions were shown to
vary significantly among cooking cycles with same
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stove, and the use of mixed fuel
(reflective of local practices) was
shown to increase plume zone BC
concentration (compared to hard
wood) by a factor of 2 to 3 across
the stoves tested.
Several studies have measured
changes in indoor concentrations
of PM2.5 (but not BC) in kitchens
in Latin America due to the
transition from a traditional
open fire to the use of a griddle
stove (known in Latin America
as a plancha stove)—a raised
wood-burning stove with a
chimney, typically designed with
a flat griddle to make tortillas.
Naeher et al. (2000) reported
reductions in 24-hour PM2.5
concentrations of over 80%, and
reported earlier measurements
that achieved reductions ranging
from 57% to 82%. (See Figure 10-10 for a picture of
an improved wood-burning stove with a chimney.)
Masera et al. found that CO and PM2.5 concentrations
in the kitchens using a so-called Patsari stove were
reduced by 66% and 67%, respectively, compared
to traditional cooking methods (Masera et al., 2007).
Johnson et al. (2007) further reported that while
Patsari stoves reduce overall particulate emissions in
homes (including net BC emissions), the BC/OC ratio
went up, making the net warming implication more
ambiguous. McCracken et al. measured personal
exposures (always less than reductions in indoor air
concentrations since individuals do not spend all
of their time in kitchens) and reported reductions
in daily average exposure to PM2.5 of over 60%
(McCracken et al., 2007).
These stoves typically cost anywhere from $8-$100
per unit, depending on the design, quality of
materials, performance, use of a chimney, use of a
metallicp/anc/ia (for making tortillas), and durability.
Certain models of these stoves have combustion
chambers that might also be used to build quality-
controlled mud stoves—the combustion chambers
themselves may cost as little as $4 to produce.
10.4.2.6 Simple Stoves
Aprovecho test results for a wide variety of
simple stoves without a rocket or other improved
combustion chamber indicated that the performance
of these stoves varies enormously, with only two
of seven tested achieving meaningful emissions
and fuels use reductions. Most achieved some fuel
savings, but increased particle emissions (MacCarty
Figure 10-10. Prakti Double-Pot Woodstove with Chimney. (Photo
courtesy of MouhsineSerrar, Prakti Design)
et al., 2010). These stoves typically cost only
$2 to $10 per unit, but may last only a few months
due to use of less durable materials and lower
quality construction.
10.4.2.7 Solar Cookers
Solar cookstoves are emissions free, and thus
the cleanest solution. However, the constraints of
current solar cookers are significant: they have
limited use in the early morning, late afternoon, or
on cloudy or rainy days; they can greatly increase
cooking time; and they are not suitable for cooking
many foods. For this reason, the potential for
current solar cookers is best thought of as part
of an integrated solution. EPA is not aware of any
example of solar cookers (which range in cost from
$20 to $75 per stove, including the pot) being
adopted at a large scale in a given region. However,
with additional advances, such as improvements in
energy storage capacity, it is conceivable that solar
stoves could be an effective tool for this field in the
future.
10.4.2.8 Behavioral and Structural Solutions
Many behavioral and structural steps can be taken
to reduce human exposures to cookstove smoke.
These include cooking outdoors, keeping children
away from cooking stoves, adding ventilation to
the kitchen, preparing fuel (drying and cutting to a
smaller size), tending stoves more carefully, lighting
stoves with improved techniques, or requiring
stoves to have chimneys. Each of these solutions
will diminish immediate human exposures to
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cookstove smoke, and are thus to be encouraged as
much as possible, though the net benefit to human
health may be tempered non-trivially by worsened
ambient air quality when use of a chimney alone is
the intervention. For purposes of this report, it is
also critical to note that some of these behavioral
and structural solutions will have little impact on
BC emissions or related climate impacts (and may
increase forcing by increasing direct emissions to
the atmosphere), while others (such as preparing
the fuel, tending stoves more carefully, and using
improved lighting techniques), may reduce climate
forcing emissions.
10.4.3 Programmatic Considerations for
Cookstove Mitigation
As this extended discussion of currently available
technologies indicates, there are a number
of promising opportunities in the cookstove
field. Advanced stoves can provide dramatic
improvements in public health, and may also offer
opportunities to reduce BC emissions. However, with
the important exception of widespread adoption of
LPG as a cooking fuel, the current scale of total stove
replacements is limited, and the number of advanced
stoves deployed as part of these programs is very
small.
There have been many efforts to bring improved
cookstoves to different parts of the world, ranging
from large-scale government efforts in both China
and India to countless small non-governmental
organization-led efforts in communities across the
globe. These efforts have had varying degrees of
success. By far the most successful effort historically
in terms of level of penetration of improved stoves
was China's National Improved Stove Program (NISP),
introduced by China's Ministry of Agriculture in
the 1980s. The NISP targeted 860 of China's 2,126
counties, and the government statistics indicate
that from 1982 to 1992,129 million improved stoves
had been installed in rural households (Graham
et al., 2005). Gradually, the Chinese government
shifted to focus on supporting stove manufacturers
(Sinton et al., 2004), and follow-on programs
increased total penetration to close to 200 million
households (Graham et al., 2005). This program was
primarily designed to reduce fuel use (to prevent
deforestation). Thus, while the use of chimneys
allowed China to lower indoor pollution somewhat,
they were not able to reduce overall air pollution and
GHG emissions (Wilkinson et al., 2009). It is not clear
what impact, if any, this effort may have had on BC
emissions.
In 1983, the government of India launched its
National Program of Improved Chuhlas (NPIC). Over
the next 17 years, the program introduced about 32
million improved biomass stoves to rural households
around the country (Barnes and Kumar, 2002). While
results varied substantially from region to region,
"A 1995-96 survey conducted by the National
Council of Applied Economic Research (NCAER)
in 18 states indicated that 71% of the cookstoves
were in working order and 60% were in use" (Sinha,
2002). Like the Chinese program, India's NPIC was
designed to lower demand for fuel wood. The
removal of indoor smoke was a secondary priority
(Partnership for Clean Indoor Air, 2005). The NPIC
has several shortcomings that limited its long-term
success, including poorly designed subsidies, poorly
designed stoves developed without user input, poor
maintenance programs, and—in most regions—no
commercial basis for sustained results (much greater
success resulted where a commercial model was
followed) (Partnership for Clean Indoor Air, 2005).
In spite of its shortcomings, India's earlier program
remains—after China's NISP—the largest cookstove
program ever implemented (Barnes and Kumar,
2002).
Ethiopia, Indonesia, Mexico, Nepal, Nigeria, and
Peru are examples of countries that have launched
national stove programs; many other countries are
actively working in this field. In December 2009, the
government of India announced that it would launch
a new National Biomass Cookstove Initiative to build
on India's earlier national program, but be based
almost entirely on a commercial business model in
close cooperation with leading manufacturers of
clean stoves and fuels in India. India will also seek
to catalyze further stove and fuel innovations, for
example via a global stove design prize.
The United States has been an active participant
in the effort to address the many health risks
associated with traditional cookstoves. At the 2002
World Summit on Sustainable Development, U.S.
EPA brought together leaders from the government,
private, academic, and nongovernmental sectors to
launch the PCIA. Through 2010, key PCIA Partners
have reported helping 6.6 million households adopt
clean cooking and heating practices, reducing
harmful exposures for more than 30 million
people. PCIA has found that succeeding with
sustainable household energy and health programs
in developing countries requires focusing on four
essential elements: meeting social and behavioral
needs of users; developing market-based solutions;
improving technology design and performance; and
monitoring impacts of interventions.
Over time, the scale and pace of cookstove
replacements have been increasing worldwide.
Based on reporting from its network of more than
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2,500,000
Figure 10-11. Number of Improved Stoves Sold by PCIA
Partners, 2003-2010. (Source: U.S. EPA)
500 partners, PCIA indicates that partners sold
2.5 million stoves in 2010 (see Figure 10-11). Based
on the latest survey results, PCIA Partners are more
than doubling their stove sales every other year.
This does not include the internal Chinese stove
market and independent manufacturers that make
and sell different versions of the so-called "Jiko"
charcoal stove across Africa. Including these sales,
the total number could be as high as 5 million to 10
million stoves per year, though there is not reliable
international data on the quality or performance
of many of these stoves, limiting the assessment of
climate and health benefits. Despite this progress,
the total impact of the cookstove replacements
to date has been small, given that the total stove
market is on the order of 500 million to 800 million
homes.
In addition to the design and fuel innovations noted
above, a number of recent developments point to
a much greater potential for making large-scale
progress in the cookstove sector. These include:
• Growth of Existing Businesses and Business
Models: An increasing number of businesses are
manufacturing and/or selling improved stoves
and fuels, utilizing a wide range of business
models. These models include non-governmental
organizations (NGOs) working to catalyze local
businesses around a common and tested stove
design (e.g., GERES Cambodia's local partners
just sold their 1.5-millionth stove); working
to develop local businesses to make and sell
artisanal stoves (e.g. GIZ's global efforts to
provide over 4 million homes with improved
stoves over the past 5 years); a local factory
selling directly (e.g., HELPS/Guatemala has
grown rapidly and sold over 100,000 stoves);
international manufactures with local distributors
(e.g., a partnership between the Aprovecho
Research Center in Oregon on design, Shengzhou
Stove Manufactures in China, and Colorado-
based EnviroFit International on sales); and major
corporations building their business in emerging
markets (e.g., Philips, Bosch-Siemens).
New Scalable Technologies: Many of the stoves
noted above represent a new suite of stove
technologies that are well designed and durable,
and for which extensive emissions testing has
been conducted. Such stoves could be mass
produced, which would improve the scalability of
these solutions (Venkataraman et al., 2010).
Carbon Financing: Cookstove businesses are
increasingly leveraging carbon financing in
both the formal and voluntary markets to
provide capital and increase public awareness.
The financing arrangements vary substantially,
but typically yield about 0.5 to 2 tons of CO2-
equivalents per stove per year for improved wood
and charcoal stoves, and up to 3 to 5 tons of CO2-
equivalents per stove per year for improved coal
stoves. Importantly, however, these credits are
based on GHG (mostly CO2) emissions reductions,
as measured by reductions in fuel use during
in-field tests. Additional work would be required
to establish credits for BC reductions. Carbon
financing is already transforming financing of
cookstove efforts into more rigorous financial
transactions with rigor and accountability for
stoves sold, stove performance in the field, and
stove utilization over time. The high transactions
costs involved in obtaining project approval also
incentivize large-scale projects and encourage
the continued use of approved stoves for many
years to generate ongoing credits. Impact
investing is a separate, but important opportunity
to bring social capital investments to this
field, and examples of this tool applied to the
cookstove field are beginning to emerge.
New Testing and Monitoring Tools: The demand
for rigorous monitoring for carbon and other
financing, research, and other needs has also led
to the development of less expensive and more
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Mitigation Approaches for Residential Heating and Cooking
effective monitoring technologies that greatly
improve our ability to measure and interpret
field results. These include relatively inexpensive
PM2.5 monitors, BC monitors, personal exposure
monitors for CO and PM2.5, portable stove
emissions testing hoods, stove use monitors, and
cell-phone based wireless monitoring tools.
In spite of this progress, achieving large-scale
adoption of clean cooking solutions will not be easy,
and many remaining barriers must be addressed.
A recent World Bank study has summarized some
of the key challenges, emphasizing the need for
a range of stoves that meet users' needs, with
demonstrated ability to reduce fuel use and indoor
smoke, while maintaining durability and safety.
The report also notes that successful programs
require functioning commercial markets in order
to reach and maintain a large scale of success
globally. Innovative financing techniques and well-
constructed monitoring and evaluation programs
were other tools highlighted as critical to success in
reaching the poor (World Bank, 2010). Other major
considerations include:
• Institutional Barriers: Such barriers include the
lack of accepted international standards for
different stove-fuel combinations, the lack of
independent stove testing facilities in market
places around the world, and the lack of health
guidelines regarding what interim targets on
what is considered a "clean" stove.
• Cost: The cost of improved stoves and fuels
alone pose a major challenge for many
households. Additional financial barriers include
tariffs and duties to import stoves, the large
investment needed to take a prototype stove
to mass production, the cost and difficulty of
developing distribution chains in target markets,
the high transactional costs of carbon financing,
and the costs of managing an inventory for a
widely fluctuating market during business start-
up. Separate financing tools are needed make
advanced stoves affordable for the poorest
populations.
• Social Barriers: Cooking solutions must be
designed to meet local cooking needs - cooking
the local food, in the timeframes needed, with
locally available fuels. Solutions for one part of
the world may not be applicable in other parts
of the world. Past "improved" stoves have not
always been designed with the needs and social
practices of end users in mind. By extensively
testing prototype stoves with users, commercial
businesses have been able to lessen these
risks. Experience indicates that a full portfolio
of solutions will be needed to meet the many
cooking needs of the developing world -
including preferences for a variety of cooking
options (just as most kitchens in the developed
world use not use a gas or electric stove, but
an oven, a microwave oven, an outdoor grill,
a toaster, and many more specialized cooking
devices).
• Global Leadership: Coordination and cross-
disciplinary leadership is needed to pursue
integrated solutions that address each of
the climate, health, gender, forestry, energy,
agricultural, and other dimensions of the
cookstove issue. In the past decade, several
new efforts have emerged that have brought
new focus to the health and climate risks of
cookstoves, and new rigor to solutions to these
risks. These include the U.S. EPA-led PCIA, the
Shell Foundation's Breathing Space program,
GIZ's HERA program, and SNV's global biogas
efforts, as well as more isolated investments by
the World Bank, USAID, and several agencies
focused on refugee camps (e.g., United Nations
High Commissioner for Refugees and World Food
Programme).
In September 2010, the United Nations Foundation
and nineteen founding partners launched the Global
Alliance for Clean Cookstoves. The Alliance is a new
public-private initiative whose mission is "to save
lives, improve livelihoods, empower women, and
combat climate change by creating a thriving global
market for clean and efficient household cooking
solutions." The Alliance will work closely with
private, non-governmental, UN and other partners
to expand efforts to address the global and local
barriers that have limited the scope of cookstove
replacements. The Alliance has set an interim goal of
having 100 million new homes adopt clean and safe
cooking solutions by 2020. Achieving this goal will
likely entail the sale through commercial distribution
channels of well over 100 million stoves in total,
with both the quantity of sales and the quality of
performance growing substantially over time (see
Figure 10-12).
The development of standards for what constitutes
a low-emitting stove is essential for ensuring
improvements in performance overtime. PCIA
and the Alliance have joined together with the
cookstove community to pursue the development
of voluntary global standards through an inclusive,
transparent process with the International Standards
Organization (ISO). The development of these
standards will proceed in parallel with growth in the
global stove market. In the early years, distribution
chains and businesses will be built around the sale
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Chapter 10
StOflAL A11JANCE FOO
OtANCOOKSTOVES
Continuous Quality Improvement of Stoves is Critical
Standards Communication and Adoption
30 -,
£, 10
Ratings System
Communication
and Adoption
1
Globally-
Recognized
Standards
Operational
Interim Stove
Rating System
Operational
Note: Numbers of different stove types in chart illustrative only
3 efficient, ultra low emission stoves
I efficient, low to moderate emission stoves
m efficient, moderate to high emission stoves
Figure 10-12. Potential Growth in the Number of Households Adopting Clean Cookstoves Globally
through 2020. The Global Alliance anticipates that the market for clean cookstoves will continue to evolve,
in parallel with development of cookstove standards. The Alliance has set a goal of 100 million clean
stoves by 2020. Numbers of different stove types in the chart are illustrative only. (Reproduced from U.N.
Foundation, 2011.)
of available, mostly mid-range stoves. As these
distribution channels are built, however, newer
advanced solutions will supplant them as the early
purchases wear out and are replaced. By the end of
the decade, it is the Alliance's goal that most stoves
sold will be of the advanced (efficient and very low
emission) variety. It is these more advanced solutions
that are likely to achieve the more significant BC
reductions, as well as the more dramatic health
benefits.
While open fires or crude stoves are not a significant
source of BC emissions in the United States, the U.S.
government has been at the forefront of the effort
to establish the Alliance and is a leading partner to
the Alliance. The U.S. Department of State is leading
Alliance diplomacy to raise the visibility of the issue
and engage new country and other partners, and
several agencies (EPA, U.S. Department of Health
and Human Services [HHS, including the National
Institutes of Health and the Centers for Disease
Control and Prevention], DOE, USAID, Overseas
Private Investment Corporation, Peace Corps,
U.S. Department of Agriculture, and the National
Oceanic and Atmospheric Administration) are
contributing substantially to the Alliance through
applied research (on technology, health, stove
testing, distribution, adoption, climate, biofuels,
forestry), financing, and distribution.
Since its launch , the Alliance has identified up to
$120 million in partner commitments, including
about $20 million for operations, over $50 million
for research, and up to $50 million in financing;
brought on more than 275 partners, including
28 country partners; catalyzed the process to
develop international consensus standards for clean
cookstoves; funded a Kenya study to assess various
cookstoves to determine the benefits for children's
health; supported regional Alliances in Asia, Latin
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Mitigation Approaches for Residential Heating and Cooking
America, and Africa to spur local businesses and
solutions; built up technical capacity of regional
stove testing centers in China, Ethiopia, and other
countries; executed comprehensive market analyses
of the clean cookstove sector in five countries;
supported the development of indoor air guidelines
by the World Health Organization; ensured inclusion
of household air pollution as a risk factor for non-
communicable diseases in the Political Declaration
for the UN General Assembly; sponsored the
Fifth Biennial PCIA Forum and two international
technical research workshops; begun integration of
the Alliance with the Partnership for Clean Indoor
Air; and worked with UN agencies to improve
collaboration among UN cookstoves and fuel
programs. The Alliance recently released a first-ever
sector-wide strategy report (called Igniting Change',
U.N. Foundation, 2011) that lays out a strategy for
universal adoption of clean cookstoves and fuels,
and its 10-year business plan is forthcoming in 2012.
Solutions on this scale are needed to resolve the
tremendous human health and environmental
burden—including the climate impacts—of
traditional cookstove use. As the above discussions
indicate, large scale success in this field may be
within reach. Substantial reductions in BC on the
order of 90% to 95% per household likely depend
on switching to cleaner fuels or advanced biomass
stoves. Such highly efficient, clean stoves help meet
multiple goals, including fuel efficiency, health
protection, low climate impacts, and reduction of
outdoor pollution (Venkataraman et al., 2010).
Currently, simple unimproved stoves dominate the
marketplace. Most current improved stove sales
are of the intermediate variety - rocket stoves or
other solutions that achieve important health11
and fuel use benefits, but will not achieve the large
health and BC benefits sought. As the Alliance
advances towards its interim goal of reaching 100
million homes, solutions will need to evolve towards
cleaner fuels and more advanced stoves to ensure
that substantial public health and BC benefits are
achieved. Additional research and innovations
are needed to bring these very clean solutions to
massive populations and to move as rapidly as
possible to achieve the health and climate benefits
that advanced stoves can bring to families and the
environment.
11 As head of the Department of Environmental Health Engineering
at Sri Ramachandra University in Chennai, India, Kalpana
Balakrishnan has said, "[These] existing improved stoves have to go
someway before they can meet a health-based standard, but they
are much, much better than the traditional stoves we have now"
(Adler, 2010).
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Chapter 11
Mitigation Approaches for Open
Biomass Burning
11.1 Summary of Key Messages
• Open biomass burning is the largest source of BC
emissions globally, affecting 340 million hectares/
year. However, total emissions of OC are seven
times higher than total BC emissions from this
sector, and better and more complete emissions
inventory data are needed to characterize the
impacts of open biomass burning and evaluate
the effectiveness of mitigation measures for
reducing BC emissions.
- Wildfires account for a large portion of BC
emissions from open biomass burning: in the
United States, for example, wildfires account
for 68% of BC emissions from open biomass
burning.
- The regions of the world responsible for the
majority of BC emissions from open biomass
burning are Africa, Asia, and South America,
with significant contributions from Russia/
Central Asia and North America. There is large
interannual and regional variability in these
emissions.
- BC emissions from open biomass burning
(predominately from widespread agricultural
burning and large wildfires occurring in the
northern latitudes) have been tied to reduced
snow and ice albedo in the Arctic.
• Certain emissions reductions techniques may yield
reductions in BC emissions from open biomass
burning; however, most of these techniques were
developed to reduce total PM2.5 emissions from
fires and there is still substantial uncertainty about
their effectiveness for reducing BC emissions
specifically, especially given diverse, site-specific
burning conditions.
• Appropriate mitigation measures depend on
the timing and location of burning, resource
management objectives, vegetation type, and
available resources. It is important to note that
fire plays an important ecological role in many
ecosystems, and prescribed burning is one of
the basic tools utilized to achieve multiple land-
management objectives in fire-dependent
ecosystems.
Expanded wildfire prevention efforts may
help to reduce BC emissions from wildfire
both domestically and globally. Successful
implementation of mitigation approaches
in world regions where biomass burning is
widespread will require training in proper burning
techniques and tools to ensure effective use of
prescribed fire.
11.2 Introduction
This chapter presents currently available information
regarding mitigation efforts and techniques that
may help reduce particle emissions from open
biomass burning (agricultural burning, prescribed
burning, and wildfires). The effectiveness of these
controls on emissions of BC and OC (including
brown carbon) requires further study. In addition,
given the importance of planned fire as a land
management tool, there are important tradeoffs
that must be considered in evaluating mitigation
options for open biomass burning.
11.3 Emissions from Open Biomass
Burning
Open biomass burning, as discussed in this
report, encompasses three main categories of
burning: agricultural burning, prescribed burning,
and wildfire.1 Table 11-1 describes each type of
open biomass burning, the land types on which
they may occur, and examples of typical resource
management objectives each burning type is
designed to achieve. In some cases, there are slight
differences in how these terms apply to domestic
and international burning practices.
The Joint Research Centre of the European
Commission estimates that 350 million hectares
(865 million acres) of land were affected by
1 Categories of contained biomass combustion, including
residential heating and cooking and industrial biomass combustion,
are addressed in previous chapters.
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Chapter 7 7
Table 11-1. Types of Open Biomass Burning. (U.S. EPA, 1998)
Type of
Burning
Agricultural
Description
The planned burning of vegetative debris from
agricultural operations. (Domestic)
The use of fire as a method of clearing land for
agricultural use or pastureland. (International)
Land Type
Forestland,
cropland,
rangeland,
grassland,
wetlands
Forestland,
rangeland,
grassland,
wetlands
Typical Resource
Management Objective(s)
Restore and/or maintain fire-dependent ecosystems;
control weeds, pests, and disease; manage lands for
endangered species; promote various vegetation
responses; reduce fuel loading to reduce catastrophic
wildfire risk; improve crop yield; control invasive
species; facilitate crop rotation; remove crop residue.
Conversion of land into cropland or pastureland.
Prescribed
The planned burning of vegetation under
controlled conditions to accomplish
predetermined natural resource management
objectives. Conducted within the limits of a
fire plan and prescription that describes the
acceptable range of weather, moisture, fuel, fire
behavior parameters, and the ignition method
to achieve the desired effects.
Forestland,
rangeland,
grassland,
wetlands
Restore and/or maintain fire-dependent ecosystems;
control weeds, invasive species, pests, and disease;
manage lands for endangered species; promote
various vegetation responses; reduce fuel loading to
reduce catastrophic wildfire risk.
Wildfire
An unplanned wildland fire (such as a
fire caused by lightning), unauthorized
human-caused fires (such as arson or acts
of carelessness by campers), or escaped
prescribed burn projects (escaped control due
to unforeseen circumstances).
Forestland,
rangeland,
grassland,
wetlands
Fire suppression or other appropriate management
response.
fire, worldwide, in 2000 (Food and Agriculture
Organization of the United Nations, 2007). However,
given the lack of an international standard for fire
terminology and the lack of consistent data reporting
and collection, it is not possible to distinguish
among the fractions of land area that were subject
to agricultural versus prescribed burning or wildfire
(Food and Agriculture Organization of the United
Nations, 2007). Generally, the mass of BC emitted
from open biomass burning will depend on the size
and duration of the fire, fuel type, fuel conditions,
fire phase, and the meteorological conditions
on the day of the burn. The emissions estimates
presented in Chapter 4 indicate that open biomass
burning represents a potentially large, though poorly
quantified portion of the U.S. BC emissions inventory.
As with the international fire emissions inventories,
available data are limited regarding the percentage
of land area affected by different types of burning.
It is also important to note that emissions of OC
are seven times higher than BC emissions from
this sector. Preliminary research suggests that the
OC fraction may be dominated by BrC, which also
absorbs light. More focused research is needed to
clarify the composition and quantity of emissions
from different types of fires.
As the estimates in Chapter 4 indicate, open
biomass burning is the largest BC source in Africa,
Central and South America, and Asia, and is one
of the largest sources of BC in Russia/Central Asia
(the former USSR) and North America. However,
there is considerable variation in the type of open
burning that dominates in different regions. Fires in
sub-Saharan Africa are primarily due to slash-and-
burn practices for clearing agricultural sites, burning
of crop residues, escaped planned burning, acts
of carelessness, and arson (Food and Agriculture
Organization of the United Nations, 2007). The
primary causes of fire in Central and South
America include large-scale conversion of moist
tropical forest to rangeland and agriculture, arson,
negligence, and hunting (Food and Agriculture
Organization of the United Nations, 2007).
Available information suggests that the majority
of fires in China and other East Asian countries are
uncontrolled wildfires, typically caused during land
conversion, or by arson and acts of carelessness
(Food and Agriculture Organization of the United
Nations, 2007). Prescribed burning is used to some
degree in China to reduce catastrophic wildfire
risk (Morgan, 2009). In India, and other South and
Southeast Asian countries, fire emissions stem from
agricultural burning, rangeland clearing, escaped
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Mitigation Approaches for Open Biomass Burning
planned burning, or acts of carelessness (Food and
Agriculture Organization of the United Nations,
2007). Agricultural burning in Kazakhstan, southern
Russia, Central and Eastern Europe is a seasonal
occurrence, typically starting at the end of April and
lasting for a few weeks (Warneke et al., 2009; Stohl
et al., 2007). Wildfires in Russia (Siberia) are primarily
caused by lightning, escaped planned burning,
or acts of carelessness (Food and Agriculture
Organization of the United Nations, 2007), and
occur from late April throughout the summer
(Warneke et al., 2009; Generoso et al., 2007). Russia
experiences many smoldering fires in drained or dry
peatlands that burn for long periods and produce
large quantities of smoke (Food and Agriculture
Organization of the United Nations, 2007). In the
Far East and southern Siberian portions of Russia,
extensive prescribed burning of the grasslands has
been used in the spring to reduce highly flammable
surface fuels (Food and Agriculture Organization of
the United Nations, 2007).
As described in Chapters 2 and 4, there is strong
evidence to suggest that emissions from fires in one
world region can significantly impact other world
regions through transport and deposition processes.
Reduced snow and ice albedo, and increased rates
of melting in the Arctic, the Himalayas, and other
snow- and ice-covered regions of the world are
major impacts of BC deposition, with implications for
freshwater resources in regions dependent on snow-
fed or glacier-fed water systems. Most of the BC that
reaches the Arctic has been traced to sources in the
Northern mid-latitudes (AMAP, 2009), with open
biomass burning as one of the largest of the sources.
A primary determinant of the downwind impact of
a large fire on snow- and ice-covered regions is the
height to which the plume rises (i.e., its injection
height). Fire plumes observed by satellite between
1978 and 2009 have shown that more dense wildfire
plumes rose to the level of the free troposphere (i.e.,
8 km), where long-range transport can occur more
readily, over North America than over Australia, or
Russia and Northeast Asia (Guan et al., 2010). This
difference has been attributed to the type of wildfire
that dominates in North America (i.e., boreal crown
fires2 that are large and very high in temperature).
In general, between 5 and 28% of the plumes from
large wildfires in North America rise into the free
troposphere (Val Martin et al., 2010).
Current emissions projections suggest that direct PM
emissions from open biomass burning will continue
to dominate global BC inventories. In addition,
several major climate change science assessments
have concluded that large, catastrophic wildfires
will likely increase in frequency over the next
several decades because of climate warming (Field,
2007; Ryan et al., 2008; Wiedinmyer and Hurteau,
2010; Littell et al., 2010). General climate warming
encourages wildfires by extending the summer
period that dries fuels and promoting easier ignition
and faster spread (Field, 2007). Earlier spring
snowmelt has led to longer growing seasons and
drought, especially at higher elevations where the
increase in wildfire activity has been greatest (Field,
2007). Increased temperature in the future will likely
extend fire seasons throughout the western United
States, with more wildfires occurring both earlier
and later than is currently typical, and will increase
the total area burned in some regions (Field, 2007).
Within Arctic regions, climate change is expected to
shift the treeline northward, with forests replacing
a significant portion of land that is currently tundra
and tundra vegetation moving into currently
unvegetated polar deserts (ACIA, 2004). Changes
in Arctic climate are also expected to increase the
frequency, severity, and duration of wildfires in
boreal forests and dry peat lands, particularly after
melting of permafrost (Schneider et al., 2007; ACIA,
2004).3 These climate-related changes in wildfire
location, duration, and frequency will affect both BC
and OC emissions.
11.4 Fire as a Resource Management
Tool
Fires play an important ecological role across the
globe, benefiting those plant and animal species
that depend upon natural fires for propagation,
habitat restoration, and reproduction. Most
North American plant communities evolved with
recurring fire and are dependent on recurring fire
for maintenance. Ecosystem fire regime analysis
includes information about the necessary fire return
interval which may vary from one to two years
for prairies, three to seven years for some long-
needle pine species, 30-50 years for species such as
California chaparral, and over 100 years for species
such as Lodgepole pine and coastal Douglas-fir.
Natural fires also reduce fuel load, unnatural
understory, and tree density, helping to reduce the
risk of catastrophic wildfires. In many parts of the
United States, historical land management practices
during the late 19th and early 20th centuries (e.g.,
fire suppression, logging, and livestock grazing)
2 Crown fires occur in the tops of trees and are spread more quickly
than ground fires. Boreal forests are generally defined as those
occurring at high northern latitudes across North America and
Eurasia, below the Arctic tundra.
3 Peat fires are also becoming more common in the lower 48 states,
as illustrated by the Evans Road Fire (2008) and Pains Bay Fire
(2011), both of which occurred in North Carolina. Peat fires have
very high emissions relative to fires involving other types of fuels.
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Chapter 7 7
have altered the natural fire regime, changed forest
structure, and led to heavy fuel accumulation
in forests. This, in turn, has increased the size of
wildfires and total area burned (Miller et al., 2009;
Noss et al., 2006; Allen et al., 2002; McKelvey and
Busse, 1996). Accumulated fuel loads will likely
continue to affect the size and frequency of large
wildfires in the coming decades.
In the United States, prescribed burning is one of the
basic tools relied upon by land owners and managers
to achieve multiple management objectives in fire-
dependent ecosystems. When one management
objective is to maintain a fire-dependent ecosystem,
the effects of fire cannot be duplicated by other
tools. Prescribed fire can also be used to reduce
heavy fuel loads, which has the benefit of helping to
prevent catastrophic wildfires.
The following section includes an outline of
strategies that can be used for conducting
prescribed and agricultural burning in a manner that
protects air quality by reducing smoke emissions,
and managing burning conditions to protect
downwind populations. In addition, the importance
of fire prevention is discussed. These methods
may also be applied with the goal of reducing
BC emissions overall, and/or the goal of reducing
downwind deposition of BC on snow and ice. As will
be discussed, the techniques listed may be more
useful in some ecosystems than in others. Further
study is needed to identify appropriate strategies to
apply under each circumstance.
11.5 Smoke Mitigation Technologies
and Approaches in the United States
Appropriate mitigation of BC from open biomass
burning depends on the type, timing, and location
of burning and must balance multiple objectives
including resource management, climate protection,
and health protection. Currently available literature is
extremely limited regarding the effectiveness of any
given mitigation practice for reducing BC emissions
from the three general types of burning. More
research is needed to better understand the efficacy,
potential unintended consequences, and cumulative
effects arising from the implementation of any
proposed mitigation techniques.
As a starting point, however, it is appropriate to
consider how approaches currently used to manage
the air quality impacts of open biomass burning
may be applicable to BC. Most U.S. domestic policies
and programs at the local, state, and federal level
focus on protecting air quality and public health
by managing smoke and minimizing PM emissions.
There are two basic approaches that are commonly
applied to manage the air quality impacts from open
biomass burning: (1) use techniques that reduce
the emissions produced for a given area; and/or (2)
redistribute the emissions through meteorological
scheduling and by sharing the airshed (Ottmar et al.,
2001).
One common approach in the United States for
limiting the impacts of open biomass burning is the
development and application of smoke management
programs. The Interim Air Quality Policy on Wildland
and Prescribed Fires (U.S. EPA, 1998)4 recognizes
the role fire plays as a resource management tool.
The policy addresses wildland and prescribed
burning managed for resource benefits on public,
tribal, and privately-owned wildlands. The policy
integrates two public policy goals: (1) to allow fire to
function, as nearly as possible, in its natural role in
maintaining healthy wildland ecosystems and, (2) to
protect public health and welfare by mitigating the
impacts of fire emissions on air quality and visibility.
The policy encourages state and tribal authorities
to adopt and implement smoke management
programs to mitigate the public health and
welfare impacts from prescribed fires and promote
communication and coordination of prescribed
burning among land owners. A smoke management
program can be extensive and detailed, or can
simply identify basic smoke management practices
for minimizing emissions and controlling impacts
from a prescribed fire.
Based on regulations, the EPA allows the use of
basic smoke management practices in lieu of smoke
management programs, where appropriate. The
Agency intends to issue guidance on the use of
basic smoke management practices in the revised
Air Quality Policy on Wildland and Prescribed Fires
when it is finalized. Basic smoke management
practices could include, among other practices,
steps to minimize air pollutant emissions during
and after the burn, evaluate dispersion conditions
to minimize exposure of sensitive populations,
and identify procedures to ensure that burners are
using basic smoke management practices. USDA
recently issued a technical document outlining some
potential basic smoke management practices.5
A smoke management program establishes a basic
framework of procedures and requirements for
4 As discussed in EPA's 2007 Final Rule on the Treatment of Data
Influenced by Exceptional Events (72 Federal Register 13560), the
Interim Air Quality Policy on Wildland and Prescribed Fires is
currently under revision.
5 See http://www.nrcs.usda.gov/Internet/FSE_DOCUMENTS/
stelprdbl046311.pdf.
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Mitigation Approaches for Open Biomass Burning
planning and managing smoke from prescribed fires.
It is typically developed by a state/tribal agency
with cooperation and participation by various
stakeholders (e.g., public/private land owners/
managers, the public). If a state/tribe determines
that a smoke management program is needed, they
may choose to develop a program using an array of
smoke management practices/emissions reduction
techniques that they believe will prevent air quality
violations and address visibility impairment. A smoke
management program can range from a purely
voluntary program to a program where prescribed
fires are regulated by a permitting authority that
analyzes meteorological conditions and air quality
considerations and authorizes burning by time of
day, fire location/size and anticipated duration. The
more-structured program may include enforceable
requirements on who may burn and when burning
may occur.
The basic elements of a smoke management
program include guidelines or requirements
regarding authorization to burn, coordination
and scheduling, and air quality assessment (U.S.
EPA, 1998). In cases where smoke management
programs are developed, these generally focus on:
(1) actions to minimize emissions (emission reduction
techniques); (2) evaluation of predicted smoke
dispersion; (3) public notification; (4) contingency
measures to reduce exposure; and (5) fire monitoring
and plume dispersion characteristics. In addition,
smoke management programs frequently lay out
guidelines or requirements for recordkeeping
and reporting; public education and awareness;
surveillance and enforcement; and program
evaluation.
In developing a smoke management program,
authorities have a number of options available
for reducing emissions (e.g., emissions reduction
techniques (ERTs)) and managing smoke that can
be applied under different circumstances. It is
important to note, however, that decisions regarding
the appropriate use of different techniques are
influenced by a number of considerations—including
but not limited to air quality impacts, water quality
impacts, Endangered Species Act requirements,
and basic resource management objectives. It is
also important to note that land managers take
safety into consideration when choosing smoke
management strategies and ERTs. The following
section provides an overview of the current practices
employed for mitigating air quality impacts.
11.5.1 Managing Smoke
Many methods for managing smoke, including
emissions reduction techniques, may offer the
benefits of reduced BC emissions and reduced
downwind impacts related to BC deposition
on snow and ice, although there are significant
uncertainties regarding transport of prescribed fire
emissions to the Arctic regions. However, there is
still substantial uncertainty about the applicability
and effectiveness of these emissions reduction
techniques for reducing BC under diverse, site-
specific burning conditions. The appropriateness of
a given mitigation practice and its effectiveness at
reducing PM2.5 and/or BC will depend on the type
of fuel being burned (e.g., crop residue or forest),
the management objectives of the burn, and the
seasonal timing and geographic location of the
burn. An additional consideration is that open
biomass burning occurs on land under various
ownership (i.e., federal, state, tribal, and private),
which affects management decisions and the types
of burning practices implemented on those lands.
Currently available literature identifies a number of
current fire management practices to address air
quality impacts of PM emissions from agricultural
and prescribed burning. These practices are listed
below.
11.5.1.1 Agricultural Burning PM Mitigation
Techniques
• Reduce the number of acres burned
- Reduce burning through conservation tillage,
soil incorporation, or collecting and hauling
crop residues to central processing sites
(WRAP, 2002).
- Apply alternate year burning which involves
alternating open field burning with various
mechanical removal techniques. The period
may involve burning every other year or every
third year (U.S. EPA, 1992).
• Increase combustion efficiency
- Use bale/stack for agricultural residue. The
bale/stack burning technique is designed
to increase the fire efficiency by stacking or
baling the fuel before burning. Burning in
piles or stacks tends to foster more complete
combustion, thereby reducing PM emissions.
This control is applicable to field burning
where the entire field would be set on fire,
and can be applied to all crop types (U.S. EPA,
2005b).
- Use propane flamers as an alternative to
open field burning.
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Chapter 7 7
- Use backing fires ("backburning"). Flaming
combustion is cleaner than smoldering
combustion. Backburning ensures more fuel is
consumed in the flaming phase (Ottmar et al.,
2001).
• Reduce fuel loadings
- Remove straw/stubble before the burn.
• Change burn timing from early spring to either
winter or summer to reduce higher impact of
BC on snow/ice. Quinn et al. (2008) suggest
that this technique may be especially important
for mitigating climate impacts in the Arctic, to
reduce springtime deposition when the snow and
icepack is large. Applicability of this technique
will be limited by the type of crop, the resource
objectives sought, and biological and operational
constraints.
• Convert Land Use
- Convert from a crop that requires burning to a
crop that does not.
- Convert land to non-agricultural use.
• Educate Farmers
- Provide training to farmers on proper burning
techniques that reduce emissions.
11.5.1.2 Prescribed Burning PM Mitigation
• Reduce the area burned
- Use mosaic burning. Landscapes often contain
a variety of fuel types that are non-continuous
and vary in fuel moisture content. Prescribed
fire prescriptions and lighting patterns can be
assigned to use this heterogeneity in fuel and
fuel moisture to mimic a natural wildfire and
create patches of unburned areas or burn only
selected fuels (Ottmar et al., 2001).
• Reduce fuel consumed (Ottmar et al., 2001)
- Burn fuel when moisture content is high.
Fuel consumption and smoldering can be
minimized by burning under conditions of
high fuel moisture of duff, litter, and large
woody fuels.
- Conduct burns before precipitation.
Scheduling a prescribed burn before a
precipitation event may limit the consumption
of large woody material, snags, stumps, and/
or organic ground matter.
Reduce fuel loadings (Ottmar et al., 2001)
- Burn outside the growing season, burn
after timber harvest, and burn frequently.
Prescribed burning at appropriate times
can help reduce the size and magnitude of
wildfires.
- Expand the use of biomass. Harvesting
and selling or trading the biomass is one
alternative to prescribed burning. Woody
biomass can be used in various industries
such as pulp and paper, methanol production,
and garden bedding. This alternative is
most applicable in areas that have large
diameter woody biomass and the biomass
is plentiful and accessible so as to make
biomass utilization economically viable.
Small-diameter biomass can be used as
posts, poles, or tree stakes. Neary and Zieroth
(2007) documented a successful USDA Forest
Service project in Arizona to remove and
sell small-diameter trees for use in small
power plants that burn wood fuel pellets.
Biomass can also be pyrolyzed to produce
biochar, a fine-grained charcoal, for use as
a soil amendment (i.e., to improve physical
properties of the soil, such as water retention,
permeability, water infiltration, drainage,
aeration and structure).6
- Use other fuel treatments such as mechanical
treatments/removal. Mechanical treatments
may be appropriate when management
objectives are to reduce fuel density to
reduce a wildfire hazard, or to remove
logging waste materials (slash) to prepare a
site for replanting or natural regeneration.
On-site chipping or crushing of woody
material, removal of slash for off-site burning
or biomass utilization, whole tree harvesting,
and yarding (pulling out) of unmerchantable
material may accomplish these goals.
Mechanical treatments are normally limited to
accessible areas, terrain that is not excessively
rough, slopes of 40% or less, sites that are
not wet, areas not designated as national
parks or wilderness, areas not protected
for threatened and endangered species,
and areas without cultural or paleological
resources.
' See http://www.biochar.org.
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Mitigation Approaches for Open Biomass Burning
- Use chemical treatments. When the
management objective is to preclude, reduce,
or remove live vegetation and/or specific plant
species from a site, chemical treatments may
be appropriate tools. However, other potential
environmental impacts caused by applying
chemicals must also be considered.
- Use animal grazers. Increasing grazing by
sheep, cattle, or goats before burning on
rangeiands and other lands can reduce grassy
or brushy fuels prior to burning, and can help
reduce burn frequency.
• Increase combustion efficiency
- Use mass-ignition techniques that produce
short-duration fires (e.g., aerial ignition).
Mass ignition can shorten the duration of the
smoldering phase and reduce the amount
of fuel consumed. However, mass-ignition
may also increase plume rise. Therefore, all
methods should be evaluated specifically for
BC.
- Use backing fires (see above).
- Burn piles or windrows. Fuels concentrated
into clean and dry piles or windrows generate
greater heat and burn more efficiently.
- Use air curtain incinerators, which are large
metal containers or pits with a powerful fan
device to force additional oxygen into the fire,
to produce a very hot and efficient fire with
very little smoke. Air curtain incinerators offer
a useful alternative to current fuel reduction
and disposal methods, providing the
benefits of producing lower smoke emissions
compared to pile or broadcast burning;
burning a greater variety, amount, and size
of materials from dead to green vegetation;
reducing fire risk; operating with fewer
restrictions in weather and burn conditions;
and containing burn area to a specific site.7
• Education for Resource Managers
- Train resource managers on proper burning
techniques to reduce emissions.
Currently available literature is extremely limited
regarding the cost of reducing BC emissions from
agricultural and prescribed fire. Many of the PM
emission reduction techniques described above
require substantial infrastructure and resource
investment (e.g., roads, machinery, etc.) or the
existence of a market for biomass utilization
products (e.g., wood pellets or biochar). The
availability of the required infrastructure, resources,
and markets will vary across the country, making the
cost of potential mitigation options highly uncertain
and dependent on the technique(s) and the site-
specific environmental conditions in which the
technique(s) are applied. A recent study (Sarofim,
2010) surveyed currently available literature
to develop rough cost estimates for the major
categories of PM emission reduction techniques
described above (i.e., increase combustion efficiency,
reduce fuel consumed, reduce fuel loadings, and
reduce the area burned).8 The authors found that
these techniques are on the whole likely to be quite
expensive for the amount of BC reduced, although
there may be potential for lower cost mitigation
approaches in locations where markets for biomass
utilization exist.
11.5.2 Fire Prevention Techniques
While wildfires are part of the natural functioning
of many ecosystems, increasing fuel loads within
the United States over the past century have made
wildfires harder to control and more expensive
to suppress. In addition, wildfires often pose a
dangerous threat to the lives and property of
civilians and firefighters. Fire prevention techniques
can be effective in helping to prevent unplanned
human-caused fires. Wildfire prevention efforts
in general can be seen as an important strategy
for limiting BC emissions both domestically and
globally. A number of studies have discussed the
timing and structure of prevention efforts to ensure
optimal effectiveness in limiting the extent and
severity of wildfires (Prestemon et al., 2010; Butry et
al., 2010a; Butry et al., 2010b).
Efforts by the U.S. Forest Service and other resource
management agencies are currently underway to
turn fire suppression programs into more proactive
fire management programs that effectively apply
fire prevention and hazardous fuels reduction
techniques, extensive public education, and law
enforcement (National Interagency Fire Center,
2011).
7 See http://www.fs.fed.us/eng/pubs/html/05511303/05511303.html.
8 The authors calculated unit emissions reductions of the various
mitigation options using emissions factors in tonnes of BC/OC per
kilogram of dry matter burned. Because these emissions factors
vary according to the particular crop/ecosystem burned and the
phase of burning (e.g., flaming or smoldering), there was a range of
values each open biomass burning source category. Sarofim et al.
(2010) used the median (when multiple data points were available)
or the midpoint (when only two data points were available) of the
range.
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Chapter 7 7
Fire prevention approaches involve a combination
of engineering, education, and enforcement.9
Education strategies often represent low-cost
approaches for preventing unplanned fires.
Such strategies must include clear planning and
communications with regard to subjects such as
fire-prone areas where access is closed or restricted;
appropriate use of campfires, smoking, and
fireworks; and managing the burning of trash and
debris. Raising public awareness through education
and outreach, including utilizing media such as
newspapers, radio, and television, is also important.
Such educational campaigns can be highly effective
in preventing unwanted fires: the U.S. Forest
Service's long-standing Smokey Bear campaign
is among the most successful fire prevention
awareness and education campaigns ever conducted
(National Wildfire Coordinating Group, 2007).
11.6 Mitigation Technologies and
Approaches Globally
As discussed in Chapter 2, a number of recent
studies have pointed to the importance of reducing
international BC emissions from open biomass
burning to alleviate effects on the Arctic, the
Himalayas, and other key snow and ice-covered
regions. Many of the mitigation techniques and
approaches described above could also be applied
internationally, and such strategies could provide
important climate benefits. However, the practical
mitigation options available on the ground in
different regions are limited for a number of
reasons. Critical barriers to implementing mitigation
measures internationally fall within three areas: (1)
weak governance (e.g., requisite laws and policies
at all levels of government to authorize and enforce
fire management practices); (2) lack of local capacity
(e.g., requisite funding, training, equipment, and
human resources to implement fire management);
and (3) lack of support infrastructure (e.g., roads and
other infrastructure to access rural areas prone to
wildfire, monitoring and early warning systems to
detect and track fires).
According to the Food and Agriculture Organization
of the United Nations (2007), many African countries
particularly in sub-Saharan Africa have no central
government fire management policy, and there
is a widespread lack of support infrastructure,
funding, equipment, and adequately trained
human resources for fire management. While most
countries in Central, South and Southeast Asia have
a government fire policy, limited funding resources
9 Additional information on each of these strategies is available on
the National Wildfire Coordinating Group's publications page at
http://www.nwcg.gov/teams/wfewt/products.htm.
restrict their ability to establish or maintain effective
fire management programs (Food and Agriculture
Organization of the United Nations, 2007).
According to Morgan (2009), the Association of
Southeast Asian Nations instituted a "zero burning"
policy in 1999, but it has been largely ineffective.
China, Japan, and South Korea have advanced fire
detection systems, including the use of remote
sensing (Morgan, 2009), but often at the local level,
villages and communities lack resources, adequate
training, and professional expertise to control
large wildfires (Food and Agriculture Organization
of the United Nations, 2007). In many countries
in South America, illegal burning even on state-
protected lands is widespread due to the absence
of enforcement and criminal penalties (Food and
Agriculture Organization of the United Nations,
2007). Russia has well-defined laws regulating forest
burning practices, but lacks strong enforcement
(Food and Agriculture Organization of the United
Nations, 2007).
Given these challenges, addressing fundamental
barriers to implementation may be just as or more
important than identifying and promoting more
technological forms of mitigation such as specific
burning techniques. Capacity-building efforts
may include building basic fire management
infrastructure, strengthening governance structures
to create and enforce fire policies, and developing
economic alternatives to slash-and-burn agriculture.
In addition, fire prevention efforts may be important
for mitigating wildfire globally. Fire prevention
education for the general public and training for
workers in the agricultural and forestry sectors in
the controlled use of fire will also be important.
There is relatively little information regarding costs
of open biomass burning mitigation internationally.
Mitigation costs will vary according to country,
and will likely be higher in developing countries
due to more extensive barriers to implementation
as described above. These costs will depend
on local environmental conditions, ecosystem
type, fire management capacity, and support
infrastructure. Costs may also vary within individual
countries, according to locality, because authority
and responsibility for fire management is often
decentralized and is left up to local or regional
authorities (Food and Agriculture Organization of
the United Nations, 2007).
To address the impact of open biomass burning
internationally, the United States has recently
initiated research efforts and other international
cooperative activities to evaluate and reduce BC
emissions from open biomass burning in and around
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Mitigation Approaches for Open Biomass Burning
the Arctic. The U.S. State Department is coordinating
a $5 million Arctic Black Carbon Initiative that will
fund a number of activities, including a project by
the U.S. Department of Agriculture (USDA) to address
biomass burning emissions in Eurasia. USDA's multi-
agency program contains the following components
(USDA, 2010):
• Research Activities: USDA scientists (led by
the U.S. Forest Service and Agricultural Research
Service) will seek to improve estimation of
emissions and transport of BC from agricultural
burning and forest fires by quantifying spatial
and temporal patterns of these emissions in
Eurasia and conducting an assessment of long-
range transport of BC from fires in Russia and
adjoining regions to the Arctic. The research will
identify meteorological conditions and potential
source locations for Arctic transport of smoke and
analyze agronomic practices in Eurasia to identify
opportunities for reduced use of agricultural
burning. Initial results from this project are shown
in section 4.5 of this report, which discusses long-
range transport of emissions. By examining the
ability of the atmosphere to transport emissions
to the Arctic, the project can identify which source
regions are most likely to contribute to emissions
transport to the Arctic. One of the benefits of
this project is that the results are independent
of source type, and therefore applicable beyond
biomass burning. Injection height is shown to be
critical (see section 4.5 for more detail).
Technical Exchange and Other Cooperative
Activities: The U.S. Forest Service and Foreign
Agricultural Service will implement technical
exchanges and cooperation between U.S. and
Russian experts on BC, agricultural burning, and
fire management. These efforts will support
training activities and the development and
implementation of innovative local-level "pilot"
programs designed to illustrate strategies and
practices that could be more broadly applied to
reduce any negative environmental impacts of
agricultural and forest fires. Key issues include
interagency cooperation on fire management,
fire budgets, and GIS and remote sensing. USDA
will also facilitate public-private partnerships to
develop local-level fire wardens and fire brigades
in Russia and outreach to farmers in Russia to
increase awareness of approaches to reduce BC
emissions from agricultural burning.
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Chapter 12
Key Black Carbon Mitigation
Opportunities and Areas for Further
Research
12.1 Summary of Key Messages
• Mitigation of BC offers a clear opportunity:
continued reductions in BC emissions can provide
significant near-term benefits for climate, public
health, and the environment.
• Effective control technologies and approaches are
available to reduce BC emissions from a number
of key source categories.
- BC emissions reductions are generally
achieved by applying technologies and
strategies to improve combustion and/or
control direct PM2.5 emissions from sources.
These and other mitigation approaches could
be utilized to achieve further BC reductions,
both in the United States and globally.
- BC mitigation solutions vary significantly by
region, and must be adapted based on the
specific needs and implementation challenges
faced by individual countries. Some source
categories, such as mobile diesel engines,
can be controlled with similar strategies
around the world. Other source categories,
such as improved stoves for residential
heating and cooking, will require tailored
solutions designed to address local needs and
challenges. Further work is needed to refine
these options to identify which might be best-
suited to particular locations or situations.
• Achieving further BC reductions, both
domestically and globally, will require adding a
specific focus on reducing direct PM2.5 emissions
to overarching fine particle control programs.
- BC reductions that have occurred to date
(largely in developed countries) are mainly
due to control programs aimed at PM2.5, not
targeted efforts to reduce BC per se. Greater
attention to BC-focused strategies has the
potential to help protect the climate (via
the BC reductions achieved through direct
PM2.5 controls) while ensuring continued
improvements in public health (via control of
direct PM2.5 in highly populated areas). Even
if such controls are more costly than controls
on secondary PM precursors, the combined
public health and climate benefits may justify
the expense.
The options identified in this report for
reducing BC emissions are consistent with
control opportunities emphasized in other
recent assessments. These represent important
mitigation opportunities for key world regions,
including the United States.
- United States: Based on current inventories
and control technologies, mobile diesel
engines (on-road and nonroad, commercial
marine, locomotives) represent the largest
potential area of BC mitigation in the United
States. Forthcoming controls on new mobile
diesel engines are expected to reduce
these emissions by 86% by 2030. Diesel
retrofit programs for in-use mobile sources
are a valuable complement to new engine
standards for reducing emissions. Other
source categories, including emissions from
stationary sources (ICI boilers, stationary
diesel engines, uncontrolled coal-fired EGUs)
and residential wood combustion (hydronic
heaters and woodstoves), also offer potential
opportunities, but on a smaller scale due
to fewer remaining emissions in these
categories, or limits on control strategies that
are cost-effective or easy to implement.
- Globally: The suite of options for reducing
BC emissions globally is broader than
those used in the United States, and will
vary by region. Key BC emissions reduction
opportunities globally include residential
cookstoves in all regions; brick kilns and
coke ovens in Asia; and mobile diesels
in all regions. While a variety of other
opportunities may exist in individual countries
or regions, these sectors account for a large
portion of BC emissions, and studies have
clearly indicated the benefits of mitigating
these sources.
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Chapter 72
- Sensitive Regions: To address impacts in
the Arctic, other assessments have identified
the transportation sector (land-based diesel
engines and Arctic shipping); residential
heating (wood-fired stoves and boilers); and
forest, grassland and agricultural burning
as primary mitigation opportunities. In the
Himalayas, studies have focused on residential
cooking; industrial sources (especially coal-
fired brick kilns); and transportation, primarily
on-road and off-road diesel engines.
A variety of other options may also be suitable
and cost-effective for reducing BC emissions,
but these can only be identified with a tailored
assessment that accounts for individual countries'
resources and needs.
- Some potential sectors of interest include
agricultural burning, oil and gas flaring, and
stationary diesel engines in the Arctic far
north.
Key remaining uncertainties include:
- Atmospheric processes affecting BC
concentrations (e.g., transport and deposition)
- Aerosol-cloud interactions (e.g., radiative and
precipitation effects)
- Climate effects of aerosol mixing state
- Emissions of BC and co-emitted pollutants
from specific regions, sources
- Warming effect of non-BC aerosols in Arctic
- Impacts of BC on snow and ice albedo
- BrC climate impacts
- Shape and magnitude of PM health impact
function
- Differential toxicity of PM components and
mixtures
- Impacts of BC on ecosystems and crops
(dimming)
Important policy-relevant research needs include:
- Continued investigation of basic microphysical
and atmospheric processes affecting BC
and other aerosol species to support the
development of improved estimates of
radiative impacts, particularly indirect effects.
- Improving global, regional, and domestic BC
emissions inventories with more laboratory
and field data on activity levels, operating
conditions, and technological configurations,
coupled with better estimation techniques for
current and future emissions.
- Focused investigations of the climate impacts
of brown carbon (BrC).
- Research on the impact of aerosols in snow-
and ice-covered regions such as the Arctic.
- Standardized definitions and improved
instrumentation and measurement
techniques for light-absorbing PM, coupled
with expanded observations.
- Continued investigation of the differential
toxicity of PM components and mixtures and
the shape and magnitude of the PM health
impact function.
- More detailed analysis and comparison of
the costs and benefits of mitigating BC from
specific types of sources in specific locations.
- Refinement of policy-driven metrics relevant
for BC and other short-lived climate forcers.
- Analysis of key uncertainties.
12.2 Introduction
Based on currently available evidence, mitigation of
BC offers a clear opportunity: continued reductions
in BC emissions can provide significant near-
term benefits for climate, public health, and the
environment. The existing literature indicates that
carefully designed programs that consider the full
air pollution mixture (including BC, OC, and other
co-pollutants) can slow near-term climate change
while simultaneously achieving lasting public health
and environmental benefits. Furthermore, currently
available control technologies and mitigation
approaches have already been shown to be effective
in reducing BC emissions, often at quite reasonable
costs. These and other mitigation approaches could
be utilized to achieve further BC reductions. In the
United States and Europe, significant reductions in
BC emissions are already expected to occur over
the coming decades as existing regulations such as
those on diesel emissions are implemented. Some of
these same approaches could help reduce emissions
in developing countries, although the source mix is
significantly different and additional approaches are
needed to address developing countries' needs.
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Key Black Carbon Mitigation Opportunities and Areas for Further Research
As discussed in Chapter 6, studies looking at global
or regional BC emissions reduction scenarios have
clearly demonstrated large potential climate and
health benefits. The framework discussed in Chapter
7 suggests that policymakers have to consider a
number of factors in designing control programs
to achieve those benefits. There is no one-size-
fits-all mitigation prescription; rather, there is a
portfolio of mitigation opportunities, all of which
have the potential to provide climate, health and
environmental benefits if applied in appropriate
locations. Chapters 8-11 illustrate the range of
sources, technologies, and control options that
policymakers can consider for purposes of BC
mitigation.
This chapter brings all of this information together
to identify paths forward on BC, based on the best
available information about remaining emissions
and cost-effective mitigation opportunities. The
options identified in this report for reducing BC
emissions are consistent with control opportunities
emphasized in other assessments, including UNEP/
WMO (2011a), LRTAP (2010), USAID (2010a), the
Arctic Council (2011), Quinn et al. (2011), and U.S. EPA
(2011b). BC mitigation solutions vary significantly by
region, and must be adapted based on the specific
needs and implementation challenges faced by
individual countries. Some source categories, such as
improved stoves for residential heating and cooking,
will require tailored solutions designed to address
local needs and challenges. Further work is needed
to refine these options to identify which might be
best-suited to particular locations or situations. The
recently completed UNEP synthesis report, Actions
for Controlling Short-Lived Climate Forcers, that
identifies regionally appropriate mitigation options
for BC can provide important information toward this
goal (see UNEP, 2011).
12.3 Controlling Black Carbon as Part
of Broader PM2 5 Mitigation Program
As discussed in Chapter 7, policymakers designing
BC mitigation strategies have a number of separate
goals (health, climate, environment) and a variety
of alternative approaches to consider. Many of the
BC emissions reductions currently being achieved
are the result of broader programs or strategies
aimed at reducing overall PM2.5, primarily for
purposes of protecting public health. This raises
a very important question, namely: are existing
PM2.5 control programs and strategies (widely
implemented) sufficient to control BC to protect
climate, or are alternative strategies and approaches
needed to guarantee BC reductions? Asked another
way: what might policymakers do differently with
regard to BC controls in order to jointly maximize
climate and public health benefts? The degree of
overlap between programs and measures driven by
public health protection vs. climate goals is of critical
importance.
In the United States and other developed countries,
PM2.5 control programs have focused very heavily
on reductions in emissions of NOX and SOX, which
contribute to secondary PM2.5 formation. There
are several reasons for this: first, NOX and SOX
emissions from utilities and other major industrial
sources contribute the bulk of PM25 mass in most
areas; second, they are the pollutants principally
responsible for regional transport of PM25, and
therefore have been the focus of national rules;
third, reductions in NOxand SOxare highly cost-
effective. Widespread reductions in NOX and SOX
emissions have been credited with substantial
decreases in ambient PM25 concentrations,
reductions in acid deposition, and dramatic
improvements in public health and the environment
(U.S. EPA, 2011c, e). In both the United States and
Europe, programs to reduce emissions of NOX and
SOX continue to serve as cornerstones of PM25
air quality improvement programs. However,
these programs have relatively little impact on
BC emissions. This is due to a difference in major
emitting sources, and also the types of controls
applied, which may not be highly effective in
reducing direct particle emissions (including BC and
OC).
The other major component of existing PM25 control
programs is the effort to reduce emissions from
mobile sources, including diesel vehicles (on-road
diesels, nonroad diesels, commercial marine, and
locomotives) but also gasoline vehicles. These
programs, which have been adopted in both the
United States and Europe, have led to stringent
controls on NOX emissions, contributing to the
overall decline in NOX and SOX emissions discussed
above. In addition, controls on mobile sources
(particularly diesels) have been highly effective in
controlling BC emissions. Mobile source control
programs are principally responsible for the
dramatic reductions in BC emissions projected to
occur in the United States and Europe by 2030.
These reductions have been tied to substantial
public health benefits (U.S. EPA, 2007). In addition,
controls on certain industrial processes have
contributed substantially to reductions in direct
PM2.5 emissions (U.S. EPA, 2004b).
As PM2.5 control programs in developed countries
have matured, the mix of sources controlled has
gradually changed, and can be expected to change
further in coming decades. In the United States,
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Chapter 72
state and local areas have already been turning to
controls on direct PM2.5 emissions as a means of
dealing with persistent nonattainment problems.
Areas such as Detroit, Michigan; St. Louis, Illinois;
Pittsburgh, Pennsylvania (Liberty-Clairton); and
Birmingham, Alabama have either included or
have been planning to include emissions controls
of direct PM2.5 in their State Implementation Plans
(SIPs) to demonstrate future compliance of the
PM NAAQS.1'2'3'4 These control strategies include
emissions controls on industrial sources such as steel
mills, coke oven batteries, and foundries, and will
reduce BC, OC and toxic metals. Because many of
these sources are also located in urban areas where
significant numbers of people reside, substantial
public health benefits are also expected from these
emissions controls (Fann et al., 2011). As discussed
in Chapter 6, EPA's health-based $ per ton benefits
estimates indicate that these reductions typically can
provide higher levels of health benefits per ton of
emissions reduction as compared to reductions from
other sources.
For developed countries, then, it appears that
control programs are gradually evolving in a
direction toward controls that achieve large BC
reductions, and that these countries are on track
to have very stringent controls on, and significant
reductions from, many source categories. Some of
these reductions will not be fully in place for several
decades, which creates opportunities for incentive
programs to accelerate the reductions. Thus, there
may be hybrid strategies the United States and other
countries can adopt that will both increase public
health benefits (via control of direct PM2.5 in highly
populated areas) and protect the climate (via the BC
reductions achieved through direct PM2.5 controls).
Even if such controls are more costly than controls
on secondary PM precursors, the extra public health
and climate benefits may justify the expense.
In developing countries, the PM2.5 challenges are
somewhat different. Overall ambient PM2.5 and
BC concentrations are generally higher than in
developed countries, and PM25 control programs are
1 Michigan Department of Environmental Quality Air Quality Division
Draft State Implementation Plan Submittal for Fine Particulate
Matter. Available online at: http://michigan.gov/documents/deq/deq-
aqd-air-aqe-sip-pm25-l-14-08_223446_7.pdf.
2 United States Steel Corporation Granite City Works and Illinois EPA
Memorandum of Understanding.
3 Revision to the Allegheny County Portion of the Pennsylvania State
Implementation Plan; Attainment Demonstration for the Liberty-
Clariton PM2.s Nonattainment Area. Available online at: http://www.
achd.net/airqual/Liberty-Clairton_PM2.5_SIP-Apr2011.pdf.
4 Alabama State Implementation Plan for PM2.s for Birmingham,
Alabama.
less advanced. For example, some megacities in Asia
have annual average PM25 concentrations above 100
ug/m3, more than six times the air quality standard
in the United States (ESMAP, 2004). Moreover,
the mix of sources is quite different, requiring
different mitigation strategies. For example, mobile
sources often have relatively little emissions control
currently. Furthermore, there are many more small
BC sources like kilns and cookstoves that could be
replaced entirely with more advanced technologies.
Coal use is very high, and significant air quality
improvements could be obtained through switching
to cleaner fuels. These changes have already
occurred in developed countries. In general, the
public health opportunities of PM25 reductions
in developing countries are large, and controls
on BC and other directly emitted PM25 could
provide significant public health improvements.
This presents a real win-win opportunity, since BC
reductions will provide benefits to climate as well as
to public health.
It is clear, therefore, that emphasizing BC reductions
within broader PM2.5 control programs might be a
beneficial strategy for health and climate in both
developed and developing countries. However,
because of the significant differences in the sources,
emissions, technologies, costs, and implementation
challenges for BC reductions in different countries
and regions, the top-tier mitigation options—
those that can reliably achieve the largest climate,
health and environmental benefits at low cost—
will vary among regions. In some cases, options
can clearly be discounted in some regions due
to low effectiveness or high costs. In other cases,
barriers to implementation or simply insufficient
information about potential effectiveness can
render certain choices less attractive. When all
these factors are considered, the number of clearly
beneficial mitigation options can be narrowed down
significantly. The remaining mitigation options
present the clearest near-term opportunities to
achieve benefits for climate and public health in
specific regions.
12.4 Key Black Carbon Mitigation
Opportunities
The control strategies described in detail in Chapters
8-11 reflect the range of existing mitigation options
for BC emissions. Based on the BC emissions
profile of key world regions, the technologies
available (including their cost and demonstrated
effectiveness), and the implementation challenges
discussed in these earlier chapters, a wide variety
of PM2.5 mitigation measures appear likely to
provide substantial benefits to human health and
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the environment. A smaller subset of measures—
those that affect BC-rich sources and/or emissions
in sensitive regions—are the most likely to
simultaneously provide climate benefits. Measures
that promise both climate and health benefits can
be considered "win-win" opportunities, and the
lower the uncertainty around those measures (with
regard to effectiveness, cost, etc.), the greater the
confidence with which policymakers can move
forward. Low-cost measures for which there is
relatively high certainty of climate as well as health
and environmental benefits can be considered "top
tier" opportunities. This section discusses potential
top-tier mitigation options for key world regions,
including the United States. A variety of other
options may also be suitable and cost-effective for
reducing BC emissions, so this discussion should
not be seen as limiting or discouraging further
exploration or adoption of other measures. Rather,
it is intended to help clarify the "low-hanging
fruit" of BC mitigation opportunities. Importantly,
these opportunities are described in general terms,
aimed at highlighting sectors and regions where
mitigation holds particular promise. Individual
countries' strategies are likely to look different based
on their application of the factors outlined in the
mitigation framework in Chapter 7. Each country will
have a variety of sector- and technology-specific
opportunities that can only be identified with a
tailored assessment.
12.4.1 U.S. Black Carbon Mitigation
Opportunities
Based on current inventories and control
technologies, mobile diesel engines represent
the largest potential area of BC mitigation in the
United States. Other source categories, stationary
sources and residential wood combustion, also offer
potential emissions reduction opportunities, but on
a smaller scale due to smaller remaining emissions in
these categories, or limits on control strategies that
are cost effective or easy to implement. Due to U.S.
air quality regulations, a number of controls on all
of these source categories are already planned over
the next decade or two. However, further reductions
may be possible, for example with more dedicated
funding for diesel retrofits on existing engines.
Chapter 6 indicated that BC reductions in the U.S.
have the potential to provide large public health
benefits domestically, and climate benefits in terms
of global average radiative forcing and impacts in the
Arctic.
• Mobile engines: On-road and Nonroad,
Commercial Marine, Locomotives: The United
States has already enacted stringent standards for
new mobile source diesel engines that are projected
to reduce BC emissions by 86% between 2005
and 2030 as the existing diesel fleet is replaced
by these new engines. Though the standards do
not dictate the use of a specific technology, the
mobile source reductions will be achieved mostly
from DPFs on new diesel engines, in conjunction
with ultra low sulfur diesel fuel. EPA has estimated
the cost of controlling PM2.5 from diesel engines
via new engine requirements at about $14,000/
ton (2010$) (including the cost of ULSD fuel).
The public health benefits of these reductions
have been estimated at $290 billion annually in
2030(2010$).
For the existing legacy in-use mobile diesel
fleet in the United States, programs such as
EPA's National Clean Diesel Campaign and the
SmartWay Transport Partnership Program can
help achieve additional emissions reductions
through retrofits on an estimated 11 million
engines. DPFs can reduce PM2.5 emissions by up
to 99%, at a cost of $8,000 to $15,000 for passive
DPFs, and $20,000 to $50,000 for active DPF
systems (2010$). However, not all engines are
good candidates for DPFs because of old age or
poor maintenance.
Stationary sources: ICI Boilers, Stationary
Diesel Engines, Uncontrolled Coal-Fired
EGUs: Stationary source BC emissions in the
United States have declined dramatically in the
last century. Remaining emissions constitute
8% of the inventory and come primarily from
coal combustion (utilities, industrial/commercial
boilers, other industrial processes) and stationary
diesel engines. Available control technologies
and strategies include direct PM2.5 reduction
technologies such as fabric filters (also known
as baghouses), electrostatic precipitators, and
DPFs. These strategies range in cost from as little
as $48 per ton PM25 to over $24,000 per ton
PM2 5 (2010$), depending on the source category.
They may also involve millions of dollars in initial
capital costs. As mentioned above, controls on
direct PM2.5 from industrial sources are being
considered by a number of states as a strategy
for addressing persistent nonattainment
problems in key areas.
Residential Wood Combustion: Hydronic
Heaters and Wood Stoves: Residential wood
combustion (RWC) represents a small portion
(3%) of the U.S. BC inventory, but mitigation
opportunities are available. In part because
seasonal use of these sources is concentrated
in northern areas, reducing emissions may
reduce deposition on snow and ice. EPA has
already established emissions standards for
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new residential wood stoves and is working to
revise and expand these standards to include
all residential wood heaters, including hydronic
heaters and furnaces, as well as stoves. Mitigation
strategies for RWC sources include providing
alternatives to wood, replacing inefficient units
(wood stoves, hydronic heaters) with newer,
cleaner units through voluntary or subsidized
change-out programs, or retrofitting existing units
to enable use of alternative fuels such as natural
gas (fireplaces). New EPA-certified wood stoves
have a cost of about $3,600 per ton PM2.5, while
gas fireplace inserts average $1,800 per ton PM25
(2010$).
12.4.2 Global Black Carbon Mitigation
Opportunities
Outside of the United States, BC emissions
are increasing in some locations as a result of
rapid economic development coupled with the
lack of serious PM2.5 mitigation measures. The
magnitude of these emissions translates into very
substantial impacts to climate, public health, and
the environment, particularly in Asia. Global BC
mitigation opportunities can be evaluated in two
ways: first, in terms of major source types within
key emitting regions (i.e., the largest reduction
opportunities on a total emissions basis), and
second, by focusing on emissions affecting sensitive
regions (i.e., the best targeted opportunities in terms
of avoiding critical impacts). It is important to note
that effective BC mitigation at the global scale will
depend on application of a variety of strategies rather
than widespread adoption of a single strategy.
In terms of major emitting sectors/regions, there are
significant differences in emissions and mitigation
potential based on the level of development of the
emitting country(s). Opportunities in developed
regions such as Europe and Russia are in many ways
similar to those in the United States: mobile diesel
engines are very important, as are emissions from
boilers, stationary diesel engines, and residential
wood combustion (though the order of importance
varies by country). In general, European countries
have emissions patterns and control programs that
are similar to the United States, though the timing
of planned emissions reductions may vary. Russia
is distinctive in that it may also have significant
mitigation potential in the area of agricultural
burning: emissions from this source category in
Russia are thought to be large and potentially very
significant for the Arctic (CATF, 2009a). However,
significant uncertainties hinder mitigation actions in
this area at present.
In the developing world, the main BC-emitting
sectors internationally are residential solid
fuel combustion (cookstoves), industry, and
transportation. Mitigation measures to reduce
emissions from each of these sectors are presently
available and feasible on wide scales. Opportunities
in Asia range from mobile diesel engines to
industrial sources (especially brick kilns) to
residential cooking on traditional stoves. This region
is of particular importance for BC mitigation, both
because it accounts for approximately 40% of global
BC emissions and because of the effect of these on
the Himalayas and the Arctic. In Africa and Latin
America, emissions come largely from open biomass
burning and residential cookstoves. While large,
these emissions may be challenging to mitigate,
particularly those from open biomass burning (see
Chapter 11).
In considering which source categories and
locations offer the largest opportunities, it is
important to remember that there are significant
differences in control opportunities between
developed and developing countries. Options that
are feasible in the United States and Europe may
not be appropriate for application in developing
countries due to cost constraints or other barriers.
For example, advanced industrial emissions control
systems may not be applicable to source types in
developing countries, or the costs of installing such
systems may be prohibitive. It is also important to
note that the opportunities for BC reductions may
change as countries develop. For example, currently
developing countries have a higher concentration of
emissions in the residential and industrial sectors,
but the growth of the mobile source sector in these
countries may lead to an increase in their overall BC
emissions and a shift in the relative importance of
specific BC-emitting sources over the next several
decades. Finally, it is important to remember that
cost, emissions and other relevant data are not
as widely available internationally, which makes
assessment of top-tier mitigation options outside
the United States and Europe more difficult.
Several groups have developed preliminary analyses
of global strategies to reduce BC emissions (Rypdal
et al., 2005; Kandlikar et al., 2009; Cofala et al., 2007;
Baron et al., 2009; UNEP/WMO, 2011; Shindell, 2012).
These assessments have generally found the largest
BC reductions at lower costs in Asia, which has high
emissions from residential cooking and heating as
well as poorly controlled transportation and small
industrial sources. Some of these assessments also
included measures to reduce biomass burning in
Africa and South America. BC strategies ranged from
improved combustion to add-on particle controls.
The actual benefits to global or regional climate
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resulting from such strategies can, however, vary
significantly with the nature of the co-pollutants
emitted from different sources, their location, and
the nature of the controls.
Based on currently available information, key
emissions reduction opportunities internationally
may include:
• Residential Cookstoves in All Regions:
Replacing traditional cookstoves, which account
for 25% of total global BC emissions, with
more advanced technologies is one of the
most promising opportunities internationally,
in large part because of the enormous public
health benefits that could result from cooking
with cleaner fuels and stoves. This strategy
may provide particularly large benefits in Asia
(including China and India), given the proximity
of these sources to both large populations
and climate-sensitive regions (the Himalayas).
Exposures to cookstove emissions currently
lead to an estimated 2 million deaths each year.
In Asia, these emissions are also linked to the
devastating impacts of ABCs, including the
disruption of the Indian monsoon. While there is
still uncertainty about BC emissions emanating
from different stove and fuel types, cookstoves
also emit CO2 and CH4, contributing to climate
effects, and a variety of aerosols and gases (SO2,
NOX, CO, OC) which contribute to adverse health
and environmental effects. Reducing cookstove
emissions helps to alleviate all of these impacts.
The magnitude of the health benefits alone
justifies the cost of cookstove replacement on
large scales.
While the world-wide stove market is
approximately 500-800 million households,
current programs likely replace only
approximately 5-10 million improved stoves per
year. Significant expansion of such programs
utilizing locally appropriate advanced stove
technologies and fuels could achieve large-scale
climate, health and environmental benefits. The
costs range from $8-$100+ per stove. Improved
cookstove technologies all face important supply,
cost, performance, usability, marketability and
other barriers that have impeded progress in the
past. However, a number of factors point to much
greater potential to achieve large-scale success in
this sector today. Along with many corporations
and world governments, the U.S. government
has made a substantial diplomatic, research, and
financing commitment to the Global Alliance
for Clean Cookstoves, which has an interim goal
of having 100 million homes adopt clean and
efficient cooking solutions by 2020.5 Continued
investment in the Alliance will help achieve this
goal with great public health benefits.
• Brick Kilns and Coke Ovens in Asia: Emissions
from stationary sources represent 20% of the
global inventory, and several recent studies have
identified brick kilns and coke ovens as two
large contributing source categories for which
new, clean technologies are available to reduce
BC emissions at reasonable cost. Brick kilns
can be replaced by cleaner vertical shaft kilns
or Hoffman kilns. Coke ovens can be designed
with modern recovery ovens or the use of end-
of-pipe control technologies. Since the ratio
of BC to other PM constituents co-emitted by
these sources (such as OC, SO2 and NOX) is high,
mitigation measures for brick kilns and coke
ovens are likely to result in both climate and
health benefits. Furthermore, the health benefits
of emissions reductions in these sectors—both
for workers and for exposed local populations—
may offset mitigation costs.
• Mobile Diesel Engines in All Regions: Mobile
sources account for approximately 18% of the
global BC inventory, and both new engine
standards and retrofits of existing engines/
vehicles could help reduce these emissions. In the
United States and Europe, mitigation programs
are already underway. Many other countries have
begun the phase-in of emissions standards and
ULSD fuel, which is a prerequisite for the proper
functioning of DPFs. However, these requirements
lag behind in some regions, as do on-the-ground
deployment of DPFs and ULSD. As a result, there
remains significant opportunity internationally
to accelerate the deployment of clean engines
and fuels. Similar to stationary sources, these
emissions are often co-located with large
populations, making them a prime target for
mitigation programs motivated by both climate
and public health concerns. Mobile source
mitigation opportunities are likely to vary among
countries. In some regions, on-road sources
such as diesel trucks may be the dominant
mobile BC emissions source, while in other
regions, non-road sources such as locomotives
may be particularly important. In general, other
countries could achieve further BC reductions
by implementing standards for nonroad diesels,
locomotive, and commercial marine vessels
(categories 1 and 2).
In terms of mitigation measures aimed at protecting
particularly sensitive regions, most of the attention
5 See http://cleancookstoves.org.
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to date has been directed at two main world regions:
the Arctic and the Himalayas. Available studies
suggest that there are considerable differences
between the prime mitigation opportunities for
these two regions.
In the Arctic, recent work by the Arctic Monitoring
and Assessment Program (AMAP) has indicated
that mitigation measures near or within the Arctic
will provide the largest benefits for the region.
This is because direct forcing and snow/ice albedo
forcing in the Arctic per unit of emissions were
shown to be highest for emissions originating
near to or within the Arctic region, with the Nordic
countries (i.e. highest latitudes) having the highest
forcings per unit of emissions, followed by Russia,
Canada, and the United States (Quinn et al., 2011,
p. 95). Importantly, studies have also found that
reductions in OC emissions will provide benefits in
the Arctic too, since OC emissions appear to exert
positive radiative forcing over snow- and ice-covered
surfaces. Similarly, sulfate aerosol, which normally
exerts a cooling influence, appears to have a much
weaker effect over snow and ice (Quinn et al., 2011,
p. 101-2). This means that a wide array of measures
aimed at reducing PM in or near the Arctic will
provide benefits to the Arctic region. In accord with
the modeling results from AMAP (Quinn et al., 2011),
the Arctic Council Task Force on SLCF has specifically
recommended that Arctic Council countries—
including the United States, Canada, Russia, Sweden,
Norway, Denmark, and Iceland—pursue additional
emissions reductions from the following sources
(Arctic Council, 2011):
• Transportation, including land-based diesel
engines and Arctic shipping: While noting that
most Arctic nations have already adopted and
begun implementing regulations for new diesel
engines, the Task Force stressed the benefits
of retrofitting or retiring older vehicles and
equipment, accelerating the phase-in of ultra
low sulfur diesel fuels, and expanding control
requirements to additional source categories, such
as marine vessels and locomotives, as appropriate.
The Task Force noted that although BC emissions
from marine shipping are currently small, they
occur in direct proximity to snow and ice and are
expected to increase in future due to declining
summer sea ice in the Arctic. As a result, the Task
Force recommended a number of measures aimed
specifically at limiting BC emissions from Arctic
shipping.
• Residential heating: Emissions inventories for
Arctic countries indicate that wood-fired stoves
and boilers are a major source of BC in the Arctic.
The Task Force recommended Arctic countries
consider implementing stringent BC emissions
standards for stoves and boilers, adopt point-
of-manufacture performance standards, and
work to develop new technologies and incentive
programs to replace old stoves and boilers.
• Forest, grassland and agricultural burning:
The Task Force identified a variety of techniques
that could be utilized to reduce emissions from
open biomass burning, a leading source of
BC in the Arctic. Available measures include
technical assistance, demonstration projects,
and financing to encourage no-burn methods,
such as conservation tillage or soil incorporation;
information campaigns to prevent accidental
wildfires and unnecessary use of fire as a land
management technique; and expansion of
resources for fire monitoring, fire management,
and fire response.
Reductions outside of the Arctic will also provide
benefits to the Arctic region. The AMAP (Quinn et
al., 2011) modeling indicates that BC emissions from
outside the Arctic Council nations are contributing
(in aggregate) the most to total radiative forcing
(warming) in the region (see Chapter 2). This
suggests that reductions in BC emissions from
regions other than the Arctic itself (60°-90°N) will
be effective in reducing Arctic warming; thus, BC
reductions anywhere—including the measures listed
earlier in this chapter—can be expected to provide
benefits specifically to the Arctic region.
For the Himalayas, most of the attention has
focused on the three source categories mentioned
above as important for mitigation globally:
residential cookstoves, brick kilns, and mobile
diesel engines. USAID (2010a) and U.S. EPA (201b)
have both identified a number of measures that
could provide substantial climate, health and
environmental benefits in this region. Specifically,
these opportunities include:
• Residential cooking: Cookstoves are the
largest source of BC emissions in Asia, and
therefore offer the largest overall emissions
reduction opportunity in the region. Furthermore,
according to USAID, "household fuel and stove
interventions ... appear to consistently achieve
the highest reduction in black carbon emissions
per unit cost. This finding holds true for all stove
and fuel interventions examined" (USAID, 2010a,
p. 3). The study found that these interventions
"yield the highest net benefits per unit
intervention cost if health benefits are included"
(p. 4) and that the GHG co-benefits are also large.
All of these factors indicate that the residential
sector is a prime opportunity for BC mitigation.
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• Industrial sources, particularly coal-fired brick
kilns: Small-scale industrial sources have been
identified as major emitters of BC, with particular
attention given to mitigation potential from
traditional brick kilns. Adopting more efficient
technologies, using cleaner fuels, and improving
operating practices may help to reduce BC
emissions from this sector. In addition, product
substitution may allow for greater efficiency and
cleaner production (Heierli and Maithel, 2008).
The largest BC emissions reductions can be
achieved through substitution of Vertical Shaft
Brick Kilns (VSBKs), Hoffmann kilns, or tunnel
kilns for traditional Bull's Trench Kilns (BTKs) (U.S.
EPA, 2011b). However, even traditional kilns can
produce lower emissions with proper fuels and
operating procedures.
• Transportation, primarily on-road and off-
road diesel engines: In Asia, the main focus
in the mobile sector is land-based vehicles,
especially long-haul trucks and buses (U.S. EPA,
2011b). USAID (2010a) identified a number of
measures aimed at mobile sources as falling
in the "next tier" of BC abatement measures
after residential cookstove interventions.
Specifically, the study pointed to the need for
more stringent emissions standards, greater use
of DPFs for new and existing (retrofitted) diesel
vehicles, and a shift toward cleaner fuels such
as compressed natural gas (CNG) or liquefied
petroleum gas (LPG). However, the study also
noted that ULSD fuel is not available in Asia
except in a few major metropolitan areas and
called for efforts to promote desulfurization in
the region (USAID, 2010a). U.S. EPA (2011b) noted
that substantial efficiency improvements and
reductions in BC emissions could be achieved
through improvements in fleet logistics, such as
eliminating "deadheading" (i.e. trips with no load).
12.4.3 Other Mitigation Options
The source categories listed above for the United
States and other world regions offer the greatest
certainty of BC reductions that are likely to be
beneficial for climate. Mitigation measures aimed at
reducing BC emissions in sectors beyond those listed
above require further investigation. The effectiveness
of controls in other BC emissions source categories
is far less certain, and the costs and implementation
barriers potentially higher. Some may be highly
effective if designed and implemented correctly.
Several studies have suggested that controlling
agricultural burning could provide important
climate benefits (Warneke et al., 2010; CATF,
2009a). Since open biomass burning is the largest
source of BC emissions globally, accounting for
over 35% of the inventory, actions to reduce
emissions from this sector have the potential to
make a significant difference. However, it is also
important to recognize that emissions of OC are
seven times higher than BC emissions from this
sector, and that a large portion of emissions come
from wildfires. In the United States, for example,
wildfires contribute about 68% of total BC emissions
from open biomass burning. Presently, data on the
percent of land area affected by different types of
burning are very limited, and little is known about
how specific measures would impact climate, both
globally and regionally. Appropriate mitigation
measures are highly dependent on a number of
variables, including timing and location of burning,
resource management objectives, vegetation type,
and available resources. The costs of mitigation
measures are uncertain and potentially high, as they
depend on various site-specific factors. Thus, despite
the significant contribution of open biomass burning
to BC emissions worldwide, further investigation
into specific options is needed to identify feasible
mitigation opportunities in appropriate locations.
Successful implementation of mitigation approaches
in world regions where biomass burning is
widespread will require training in proper burning
techniques and tools to ensure effective and
appropriate use of prescribed fire.
Another sector that has drawn attention, especially
with regard to potential impacts in the Arctic, is
oil and gas flaring. Despite higher uncertainty
about the magnitude and impact of BC emissions
from flaring, the Arctic Council Task Force on
SLCF recommended that this source be evaluated
carefully due to the expansion of oil and gas
extraction activities in the Arctic region (Arctic
Council, 2011).
Other categories of interest include stationary
diesel engines in the Arctic far North. These sources
are located in proximity to snow and ice, are often
very expensive to operate due to fuel transport
requirements, and may adversely impact the health
of indigenous communities.
12.5 Key Policy-Relevant Scientific
Uncertainties
Many uncertainties are associated with
understanding the effects of BC emissions on air
quality, climate, health, and the environment. Some
of these may be important and large enough to
influence BC mitigation decisions - including the
decision about whether to invest in BC mitigation
at all as well as the selection of specific mitigation
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Chapter 12
More likely
to be
important
forBC
mitigation
decisions
Less likely
to be
important
forBC
mitigation
decisions
• Emissions of BC and co-emitted
pollutants from specific regions,
sources
• Aerosol-cloud interactions (e.g.,
radiative and precipitation
effects)
• Climate effects of aerosol mixing
state
Warming effect of non-BC
aerosols in Arctic
BrC climate impacts
Differential toxicity of PM
components and mixtures
Impacts of BC on ecosystems
and crops (dimming)
• Impacts of BC on snow and ice
albedo
• Shape and magnitude of PM
health impact function
Atmospheric processes
affecting BC concentrations
(e.g., transport and deposition)
as well as the climate effects
of different aerosol mixing
states (e.g., possible enhanced
absorption by BC when coated
by sulfate), are also highly
uncertain. Additional research
in these areas has the potential
to shift the current assessment
of the climate benefits of BC
mitigation in important ways.
Less certain
More certain
Key
Uncertainties related to Emissions and Concentrations
Uncertainties related to Climate Effects
Uncertainties related to Health and Environmental Effects
Figure 12-1. Key Policy-Relevant Scientific Uncertainties Related to BC.
(Source: U.S. EPA)
strategies. Other uncertainties may be important
scientific questions, but would be unlikely to
influence BC mitigation decisions.
Figure 12-1 roughly groups key uncertainties
identified in this report by level of uncertainty and
likelihood to affect BC mitigation decisions. Each
of these uncertainties pertains to an important
aspect of BC - emissions and concentrations,
climate effects, or health and environmental effects.
Uncertainties categorized as "less certain" and "more
likely to be important for BC mitigation decisions"
are those for which new information can change our
understanding of the direction and effectiveness
of BC mitigation in terms of climate, health, and
environmental benefits. These are the uncertainties
that are the highest priorities for future research.
For example, understanding the magnitude of BC
emissions from specific regions and sectors, along
with co-emitted pollutants (e.g. OC and BrC), is
critical to estimating the net warming or cooling
effect of that source. New information showing
a larger fraction of warming agents among the
emitted mixture from a particular source may
make mitigation more attractive from a climate
perspective; similarly, a smaller relative contribution
of warming agents from a source could make it
a less attractive target for mitigation. Impacts of
BC and other aerosols on clouds, both in terms
of radiative forcing and impacts on precipitation,
Uncertainties categorized
as "less certain" and "less
likely to be important for BC
mitigation decisions" are those
that are not well-understood
but for which new information
is unlikely to fundamentally
change our understanding
of BC mitigation impacts. In
some cases, the outcome of
new information may be to
change our understanding of
the magnitude of BC mitigation
impacts of climate, health, and
the environment, but not the
direction of these benefits. For
example, new information may
improve understanding of the effects of non-BC
aerosols (e.g., OC, BrC, sulfate) in the Arctic, both in
terms of impacts on snow and ice albedo, as well as
the pace of melting. Since all particles are generally
darker than snow and ice, additional information
could change the magnitude of estimated BC
mitigation benefits in the Arctic. Similarly, if
a broader range of aerosols are confirmed to
contribute to warming in the Arctic, it would
strengthen evidence that mitigating BC and related
co-emitted pollutants could provide significant
benefits to this region; however, such evidence
would not necessarily change the highest priority
mitigation measures aimed at BC-rich sources.
Although differential toxicity of PM2.5 components
and mixtures is critical for understanding the
magnitude of the health benefits achieved by
controlling PM2.5, the current literature points
to health effects of BC that are similar to those
of other PM2.5 components. Since it is very likely
that reductions in BC will be found to be health
beneficial, new information on differential toxicity is
less likely to affect BC mitigation decisions.
Uncertainties categorized as "more certain" are still
significant, even if there is relatively more evidence
on these topics as compared to those categorized
as "less certain." For example, the impacts of BC
on snow and ice are characterized by a small but
generally consistent body of literature. There is a
246
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Key Black Carbon Mitigation Opportunities and Areas for Further Research
strong need for additional research in this area, even
though currently available studies already suggest
BC mitigation efforts will provide benefits in snow/
ice covered areas. In addition, there remains some
uncertainty regarding the full extent of PM2.5 health
effects, especially in the developing world. The health
effects of PM2.5 are generally well characterized as a
result of numerous major long-term epidemiological
studies conducted in developed countries and a
number of short-term studies conducted in locations
around the world. Thus the health effects of PM2.5
are generally well-known and provide a strong basis
for controlling emissions of PM precursors and
components, including BC. However, the magnitude
and shape of the concentration-response function in
developing countries where exposure is significantly
higher remains under investigation.
12.6 High Priority Research Needs for
Black Carbon
There are a number of high priority research
topics that could help advance efforts to control
emissions from BC sources, clarify the impacts and
benefits of BC mitigation, and reduce key remaining
uncertainties, particularly those summarized in
section 12.5 and in Figure 12-1. Some of these
research topics pertain to key scientific aspects of
BC, including composition, morphology, fate, and
transport. Other important areas for further research
relate directly to information and tools necessary for
policymaking about BC. Research on these topics
would allow for more informed policy decisions
regarding mitigation of emissions of BC and related
species. Based on the information reviewed for this
report, EPA concludes that priority should be given
to research in the following areas:
1. Continued investigation of basic microphysical
and atmospheric processes affecting BC
and other aerosol species to support the
development of improved estimates of
radiative impacts, particularly indirect effects.
Some of the basic microphysical and atmospheric
processes that BC and other aerosol species
undergo are not very well understood. This includes
the mixing of BC with other aerosol species, the
atmospheric aging of BC and how aging affects BC's
climatic and health impacts, and interactions with
cloud droplets and the hydrologic cycle in general.
Incomplete understanding of the climate effects of
aerosol mixing state and of aerosol-cloud Interactions
limits the scientific community's ability to model BC
in the atmosphere and estimate its impacts. Few
global models are now able to resolve the cloud
microphysics which are of importance in determining
climatic effects. Direct radiative forcing from BC is
clearly positive and results in warming. However,
early results indicate BC emissions lead to a net
positive cloud absorption effect but both positive
(warming) and negative (cooling) semi-direct and
negative indirect effects. The net result may be
negative enough to offset some of the warming
due to the direct effects of BC. The net effect of BC
on cloud absorption, semi-direct and indirect cloud
feedbacks depends on many factors, among them
aerosol hygroscopicity, absorptivity, and number
concentration relative to background particles.
Improvements in our understanding of these basic
properties through controlled experiments and
atmospheric observations could improve climate
models, and could also inform ongoing efforts to
investigate the health effects of PM constituents,
including BC.
2. Improving global, regional, and domestic BC
emissions inventories with more laboratory
and field data on activity levels, operating
conditions, and technological configurations,
coupled with better estimation techniques for
current and future emissions.
Given the diversity and ubiquity of sources of
BC, accurately measuring and tracking emissions
of BC and Its co-pollutants from specific source
categories is a very difficult undertaking. Emissions
inventories in the United States and other developed
countries account for most source categories, but
considerable uncertainty remains. More information
on both emission factors and usage would be
helpful. In particular, emissions from key industrial
sources, flaring, residential heating, and open
biomass burning remain poorly characterized.
In general, mobile source emissions are among
the best characterized (especially in developed
countries), but improved information is still needed
for some sectors, such as nonroad mobile sources
(aircraft, locomotives, ocean-going vessels); newer
technology on-road diesel/gasoline vehicles; high-
emitting vehicles; and vehicles operating at low
temperatures.
Uncertainties are larger for BC inventories in
developing countries and globally. For these
inventories, priorities include better characterization
of emissions from residential cookstoves, in-use
mobile sources, small fires, smaller industrial sources
such as brick kilns, and flaring emissions. For sources
such as cookstoves, improved characterization
depends critically on field-based measurements
of emissions from in-use sources. In addition,
usage patterns need to be reviewed to ensure that
appropriate "activity" levels are applied to emission
factors to arrive at final emissions estimates. Finally,
Report to Congress on Black Carbon 247
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Chapter 72
fuller incorporation of regional inventories into
global inventories could improve country- and
region- specific emissions estimates.
Quantifying and reducing the uncertainties in
global, regional, and domestic emissions inventories
requires collecting source-specific emissions data
from lab or field-based measurements and gathering
information on activity levels, operating conditions,
and technological configurations. For hard-to reach
areas, improvements in estimation techniques could
significantly improve global and regional inventories.
Systematic collection and sharing of emissions data
and meta-data is important for both scientific and
policy purposes.
3. Focused investigations of the climate impacts
of brown carbon (BrC).
The role of BrC is important for determining the
potential climate benefit of mitigating sources
with high OC emissions such as biomass burning.
BrC may also play a stronger role in the Arctic
and snow- and ice-covered regions, through
deposition on ice and snow. Several aspects of
the BrC issue warrant research in order to better
characterize the climate impacts of BrC. First, multi-
wavelength measurements are needed to separate
and characterize BrC and BC. Reporting column
data by wavelength may aid model-observation
comparison, as BC and BrC differ in terms of peak
absorption. Second, BrC should be incorporated into
climate models, and the impact of BrC on net forcing
estimated independently from the impact of BC.
Atmospheric observations of BrC and experimental
methodologies for determining BrC emissions are
also needed, in conjunction with the improvements
in BrC measurements and expanded observations,
particularly in ice- and snow-covered regions
(described further below).
4. Research on the impact of aerosols in snow-
and ice-covered regions such as the Arctic.
Clarifying the role of BC, BrC, and other aerosols
in sensitive regions such as the Arctic and the
Himalayas is critically important for proper
characterization of the role of different species in
climate change in these regions. Additional work
on BC impacts should be supplemented with
expanded research on non-BC aerosols, including
suifate, nitrate, and OC (including BrC). An expanded
observational record coupled with more thorough
treatment of different aerosol species in climate
models (both described elsewhere in this section)
could help clarify the contribution of different
species to net warming or cooling in the Arctic
and could provide insight into effective mitigation
measures. Key topics include the impact of BC on
snow and ice albedo and the warming impact of non-
BC aerosols in the Arctic.
5. Standardized definitions and improved
instrumentation and measurement techniques
for light-absorbing PM, coupled with
expanded observations.
In order to accurately assess the impacts of BC
(and co-pollutant) emissions, it is essential to have
a clear understanding of the optical properties of
atmospheric aerosols and be able to trace those
to emissions from specific sources. Precise and
consistent definitions and measurements of BC
and other carbonaceous aerosols would ensure
more accurate assessment of BC emissions,
climate and public health impacts, and mitigation
options. Additional research is needed to improve
instrumentation and measurement techniques to
quantify accurately the light-absorption properties
of BC, BrC, and other aerosols; to harmonize
measurement and reference methods to standardize
definitions of BC, BrC and other compounds; and
to refine measurement techniques to collect more
data on light-absorption capacity of emissions from
specific sources. Additional representative source
measurements to better characterize BC emissions
by source category, fuel type and combustion
conditions can help improve emissions inventories
and reduce modeling uncertainties.
It is equally important to expand the observational
record for BC, including observations of atmospheric
concentrations of BC and BC deposition. Existing
measurements of BC are sparse in both spatial
and temporal coverage, even in countries with
more advanced monitoring programs such as
the United States. An expanded observational
record based on standardized measurement
techniques and instruments could provide important
information about BC transport, vertical distribution,
atmospheric interactions, and deposition.
Observations of BC and BrC deposition are
important for furthering understanding of the role
of deposition on snow and ice. Such data could be
used to inform climate models and verify impacts.
6. Continued investigation of the differential
toxicity of PM components and mixtures and
the shape and magnitude of the PM health
impact function.
A great deal of research on the health impacts
of PM2.5 and specific PM components has been
conducted over the past 15 years, and these topics
have already been identified as priorities by EPA
in the context of its periodic reviews of the U.S.
248 Report to Congress on Black Carbon
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Key Black Carbon Mitigation Opportunities and Areas for Further Research
national ambient air quality standards for PM. While
the scientific record is robust in many respects,
there are still important unanswered questions
about the relative toxicity of different constituents.
Also, research continues to inform the overall
understanding of the magnitude and nature of PM
health impacts, including more precise quantitative
information about the relationship between indoor
and ambient concentrations and health impacts.
Continued investment in this research is important,
and there is a particular need for more studies in
developing countries.
7. More detailed analysis and comparison of
the costs and benefits of mitigating BC from
specific types of sources in specific locations.
More work is needed to link BC sources in specific
regions to climate, health and environmental
impacts, all the way through the causal chain.
Improved characterization of BC control strategies,
their costs, and their net impact on radiative
forcing, as influenced by location, will help ensure
maximum climate benefits. This depends in large
part on improved emissions characterization and
measurement, as described above, but also more
refined modeling techniques capable of evaluating
regional or local scale impacts. Greater attention
should be paid to the location of the proposed
change in emissions, especially for near-Arctic
or near-Himalayan emissions. Similarly, the links
between non-radiative impacts of BC, such changes
in rain, snow, and water resources, and specific
source classes or regions have not yet been well
established; the ability to relate non-radiative effects
to aerosol (and precursor) emissions needs further
development.
Health impacts will also vary by source type and
location. As discussed above, more work on PM
components is need and this can support further
analysis to help identify emissions reductions
with maximum co-benefits for public health and
climate. While the relationship between BC and
visibility is relatively well understood, the impacts
of BC dimming on other welfare outcomes, such
as ecosystem and crop health, have not been well
quantified. Research on the human health and
environmental consequences of reductions in
specific sectors, coupled with development of
appropriate metrics (#8, below) would allow for
comparison between source sectors and regions and
improved ability to develop and meet policy goals.
8. Refinement of policy-driven metrics relevant
for BC and other short-lived climate forcers.
Some of the fundamental assumptions that go into
the calculation of policy-relevant climate metrics for
long-lived GHGs do not apply to BC; therefore, it is
difficult to apply metrics developed for GHGs to BC
and other short-lived forcers. Though "alternative"
metrics have been proposed for BC, none is yet
widely utilized. Appropriately tailored metrics for BC
are needed in order to quantify and communicate
BC's impacts and properly characterize the costs and
benefits of BC mitigation. Improved metrics could
incorporate non-radiative impacts of BC, such as
impacts on precipitation. Similarly, given BC's (and
other aerosols') direct impacts on human health,
health outcomes could also be incorporated into
such a metric. Developing methods to quantify the
benefits of BC mitigation on both climate and health
would encourage policy decisions that factor in
climate and health considerations simultaneously,
within a unified framework. Analysis is also needed
to examine how utilizing alternative metrics would
affect policy priorities and preferred mitigation
options.
9. Analysis of key uncertainties.
Systematic analysis of key remaining uncertainties
and technical gaps regarding BC could help
prioritize future research and investment by
clarifying which of these factors exert the largest
influence on modeled outcomes. Such analysis
would involve both: (1) Model intercomparison
of BC radiative forcing and climate impacts
between global and regional models, along with
comparisons to ambient measurements (including
remote sensing and tracer-based analyses); and
(2) Sensitivity analysis of the factors influencing
models' representation of (a) the net effect of a
given mitigation measure, considering all co-emitted
pollutants; and (b) the overall global and regional
contribution of BC and OC to radiative forcing and
temperature change.
Global and regional models give a different range
of predictions of BC's RF and climate impacts due to
different model configurations, parameterizations
and assumptions (model resolution, chemical and
physical processes, aging/mixing, deposition, etc.).
A comparison of model results and diagnostic
analysis will reduce remaining uncertainties
regarding BC impacts. Additional sensitivity analysis
should consider the importance of: (1) emissions
inventories utilized (including type and magnitude of
co-emissions represented); (2) model representation
of transport and vertical distribution of emissions,
aging and mixing of particles, radiative properties
of particles, and particle interactions with clouds
and snow; and (3) the use of observational data to
constrain model results and emissions estimates.
Report to Congress on Black Carbon 249
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Appendix 1
Ambient and Emissions
Measurement of Black Carbon
A1.1 Introduction
Measurements of BC and other PM constituents are
critical to understanding the climate impacts of these
substances, as well as evaluating human health and
environmental effects. These measurements serve
as important inputs to air quality forecasting and
climate models, source apportionment models, and
emissions inventories. Deposition measurements are
also needed to judge impacts on snow and ice.
Observational data for BC comes from two main
sources: ambient measurements and source-based
emissions measurements. These measurements
involve both sample collection and sample analysis
procedures, with each step having important impacts
on reported measurements. Most estimates of BC
are based on thermal-optical and filter-based optical
techniques, which classify the measured quantity
as apparent elemental carbon (ECa) and apparent
black carbon (BCa). While the terms "black carbon"
and "elemental carbon" are frequently used as labels
for quantities produced, the addition of the term
"apparent" clarifies that these are considered to
be estimates of BC concentrations. This appendix
describes the most common sample analysis
methods (thermal-optical and optical), the types of
instruments that can be used for these methods, and
key limitations in current measurement methods,
approaches, and instruments. This appendix
also describes the key sources of ECa and BCa
measurement data in the United States, in terms
of the types of ambient data collected and the
information gathered from testing of both stationary
and mobile sources. Next, this appendix describes
key applications of source-testing data, particularly
for constructing U.S. emissions inventories. Data
from other countries is reported where available and
applicable.
A1.2 Ambient Black Carbon
Measurements
BC mass concentration estimates are routinely
measured at ground-level in the ambient air or
in deposited materials, but can also be taken
in aircraft and on remote sampling platforms.
Globally, a significant amount of ambient data
has been compiled from the following types of
measurements:
• Ground-based ambient air measurements a re
taken in near real time using field analyzers or
obtained in a laboratory following collection of
PM onto a filter. This is by far the largest source
of observational data on BCa and ECa. Details on
some of the key ambient air monitoring networks
producing these data are described in Table Al-1.
• Ice core measurements of BCa and ECa have been
conducted in glaciers around the world, providing
a historical record of BC concentrations.
• Surface snow measurements have been conducted
to quantify recent BC in snow based on BCa and
ECa concentrations in locations around the world.
Snow data is much more limited in spatial and
temporal coverage in comparison to ambient
monitoring.
The concentration of carbon in PM is regularly
measured using methods based on the chemical,
physical, and light absorption properties of the
particles. The chemical and physical properties of
carbonaceous PM vary in terms of both refractivity
(the inertness of the carbon at high temperatures)
and light absorption. Each carbon measurement
technique provides unique information about
these properties. All current analysis methods are
operationally defined, meaning that there is no
universally accepted standard measurement. When
developing these methods and operational criteria,
some scientists use PM's optical properties or light-
absorbing characteristics (optical or light absorption
methods), some use its thermal and chemical
stability (thermal-optical methods), while others use
its morphology or microstructure or nanostructure
(microscopy methods). One major class of methods,
thermal or thermal-optical techniques, distinguishes
refractory and non-refractory carbon as ECa and
OCa, respectively (Figure Al-1). The second major
class of methods, optical methods, quantifies the
light absorbing component of particles as BCa,
Report to Congress on Black Carbon
251
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l/l
Sj
Table A1-1. Global Monitoring Activities. (Source: U.S. EPA)
Worldwide Air Monitoring Networksa'bfor Black Carbon
Network C?unW Years of Data . .BC; Number of Measurement Method Location of Information and/
Agency Indicator Sites or Data
ESRL/GMD Aerosol Network
Baseline Stations
Regional Stations
Mobile/Cooperative Platforms
World Data Centre for Aerosols
Nepal Climate Observatory-Pyramid
(NCO-P)
CSN/STN— PM2.5 Speciation Trends
Network1
IMPROVE — Interagency Monitoring
of Protected Visual Environments
ARIES /SEARCH— Aerosol Research
Inhalation Epidemiology Study/
Southeastern Aerosol Research and
Characterization Study experiment
NAPS— National Air Pollution
Surveillance Network
CAPMoN — Canadian Airand
Precipitation Monitoring Network
European Monitoring and
Evaluation Program (EMEP)
European Supersites for
Atmospheric Aerosol Research
(EUSAAR)
China Atmosphere Watch Network
(CAWNET)
United States/
NOAA
Global
Atmospheric
Watch
Nepal
United States
/EPA
United States
/NPS
United States/
EPRI/SC
Canada
Canada
Norwegian
Institute for Air
Research
European
Union
Chinese
Meteorological
Administration
1957-Present
1974-Present
2006-2008
1999-Present
1988-Present
1992-Present
2003-Present
2002-Present
2002-2003
2006-Present
1999-Present
BCaand/
or Aerosol
optical
properties
BCa and/or
Aerosol
optical
properties
BCa
ECa
ECa
ECa
ECa
ECa
ECa
BCa / ECa
ECa
4 Rural
3 Rural
15 Rural
-16 Rural
1 Rural
-200 urban
110 rural
(plus -67
protocol
sites)
5 Urban
3 Rural
4 rural
13 urban
29 Rural
2 Urban
12 Rural
20 Rural
6 Urban
12 Rural
Aerosol Monitoring System -
Aethalometers, Particle Soot/
Absorption Photometers,
Nephelometers, etc.
Aerosols - Light Absorption/
EBC, AOD, Light scattering
& back scattering. Size
distribution
MAAP
Thermal Optical Transmittance
Thermal Optical Reflectance
Thermal Optical Reflectance
R&PPartisol-Plus2025
R&P Partisol Model 2300
R&P Partisol Model 2300 PM2.5
Speciation Sampler
Thermal Optical Tranmittance -
Sunset Lab
Aerosol properties including -
absorption, scattering, AOD
Aethalometer/ Sunset Lab
Thermal Optical Reflectance
http://www.esrl.noaa.gov/gmd/aero/
index.html
h ttp://wdca.jrc. it/
http://gaw.tropos.de/gaw_program.
html
http://www.atmos-chem-phys-
discuss.net/10/8379/2010/acpd-10-
8379-2010.pdf
h ttp://www. epa.gov/ttnam til/
specgen.html
http://vista.cira.colostate.edu/
IMPROVE/
http://www.atmospheric-research.
com/studies/SEARCH/index.html
http://www..gc.ca/rnspa-naps/
Default.asp?lang=En&n=5COD33CF-l
http://www.msc.ec.gc.ca/capmon/
particulate_general_e.cfm
http://www.atmos-chem-phys.
net/7/5711/2007/acp-7-5711-2007.pdf
http://www.eusaar.net/files/overview/
in fras true tures.cfm
http://www.agu.org/journals/jd/jd08
14/2007JD009525/2007JD009525.pdf
3
a.
5'
-------
Worldwide Air Monitoring Networksa'bfor Black Carbon
Network
Multiple Independent Sites -two
groups by pollutant (BC & ECa) by
Vignatietal. (2010)
Country/
Agency
Multiple
Agencies
Years of Data
Various
periods
1976-2002
BC-
Indicator
BCa
ECa
Number of
Sites
11 Rural
7 Rural
Measurement Method
Various
Location of Information and/
or Data
http://www.atmos-chem-phys.
net/10/2595/2010/acp-l 0-2595-2010.
pdf
i The emphasis is on surface-based continuous air monitoring networks. Some networks listed separately may also serve as subcomponents of other larger listed networks; as a result,
some double counting of the number of individual monitors is likely.
> The information on some networks is sketchy. It is frequently unclear (1) when the network actually started up and whether all monitors were operating at that time (or were added over
time), (2) whether the pollutant measured is measured as BCa, ECa or some other surrogate for BC, (3) what the definition of urban/rural is for a given network and the exact numbers of
urban/rural monitors, and (4) what the exact nature of the measurement method is and whether it applies to all or just some sites.
Collocated at CSN sites for the period 2009 to present, there are -40 Aethalometers for measuring BC and 5 Sunset Laboratory Carbon Aerosol monitors for ECa.
n
o
3
O
3
2
S"
n
=s-
n
Q
—i
O-
O
3
Sj
l/l
Light-Absorption Classification
More
light-absorbing
Light-
absorbing
carbon
(LAC)
Black
carbon
(BCJ
Brown
carbon
(BrC)
Less
light-absorbing
Thermal-Optical Classification
More refractory
Elemental
carbon
Organic
carbon
Less refractory
* Measurement technique-specific split point
Figure A1-1. Measurement of the Carbonaceous Components of Particles. Black carbon
and other types of light-absorbing materials can be characterized by measuring their
specific light-absorbing properties, as seen on the left side of the figure (BCa/BrC/LAC).
This contrasts with other approaches to characterizing particles based on measurements
of the refractory nature of the material (inertness at high temperatures), as seen on the
right-hand side of the figure (ECa and OCa). (This figure is also as shown as Figure 5-1.)
(Source: U.S. EPA)
3
a
3
a.
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3
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3
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-t>
co
a"
n
sr
n
a
5-
o
3
-------
Appendix 7
which can be used to estimate BC concentrations
and can also indicate the existence of components
that absorb in the near-UV (i.e., brown carbon, BrC).
Light absorbing carbon (LAC) is a term used for light
absorbing substances in the atmosphere, which
includes soot and its components, BC and BrC. There
is a lack of consensus and standardization regarding
the operational criteria used, calibration materials
used, and defining characteristics or properties of
the BC measured. The methods used to measure ECa
and BCa require standardization and re-evaluation
for climate and regulatory uses.
A1.2.1 Thermal-Optical Methods, ECa
As noted in Chapter 5, thermal-optical methods
are by far the most commonly used. Since 1982,
thermal-optical analysis methods have been
applied to measure the ECa and OCa component
of ambient and source aerosols (Huntzicker et al.,
1982; NIOSH, 1999; Birch and Cary, 1996; Chow et
al., 1993; Peterson and Richards, 2002; Chow et al.,
2007). PM collected on a filter is heated to isolate
the refractory and non-refractory carbon. Laser
correction measurements help prevent charred
organic materials from being misinterpreted as ECa.
Thermal optical-reflectance (TOR) methods use
reflectance for char correction and separation of ECa
from OCa, while thermal-optical transmittance (TOT)
uses transmittance. Long-standing reliance on these
methods—which measure ECa, rather than BC—has
resulted in an extensive observational record based
on ECa and OCa splits, and the frequent substitution
of ECa data for BCa data, since availability of the
latter is limited. In addition to laboratory-based
thermal-optical methods for ECa, semi-continuous or
near real-time thermal-optical methods for ECa and
OCa are commercially available. The semi-continuous
analyzer provides hourly in-field measurements of
ECa and OCa. This semi-continuous analyzer also
provides a measure of light absorbing or optical BCa.
A1.2.2 Light-Absorption Measurements, BCa
Currently, light-absorption or "optical" measurements
of BCa are not consistently deployed in routine
monitoring programs in the United States. The one
program area in which light-absorption methods are
used is in assessing visibility impairment in national
parks and wilderness areas via the Interagency
Monitoring of Protected Visual Environments
(IMPROVE) program. To date, optical methods have
not been widely used in urban monitoring networks.
However, such approaches are commercially
available and could be more widely deployed.
These approaches fall into two general categories—
optically absorbing and incandescent (thermal
emission of light) measurement. Relative to the
incandescence techniques, optical techniques
for BCa are in wider use. A listing of a variety
of commercially available instruments used for
monitoring ambient or source concentrations of BCa
and the wavelength selected for measurement is
provided in the Table Al-2.
Modern light-absorbing techniques rely on passing
a laser beam at a specific wavelength through a
particle sample, either in an air volume or deposited
onto a filter, and observing how much light is
absorbed by the particles. BCa is typically measured
over the green to infrared wavelengths, where it
absorbs more strongly than other LAC. BrC may also
absorb light at shorter wavelengths (near-UV and
UV). Many BCa instruments can measure at multiple
wavelengths, sometimes simultaneously depending
on the exact instrument configuration. This provides
information about components that absorb light
over different parts of the UV/Visible spectrum.
Thus, these instruments may be used to distinguish
between BCa and BrC; however, in many cases
researchers have not been careful to distinguish
how much of the measured light-absorbing carbon
falls into each category. In order to convert light
absorption to a BC mass concentration, a mass
absorption coefficient or similar conversion factor is
used. The conversion factor is based on experiments
that simultaneously measure light absorption at
a specific wavelength and BC mass (either as ECa
from ambient measurements or particle mass from
soot generation experiments). It is recommended
that light absorption be reported in the original
units of absorption along with any mass absorption
coefficients or conversion factors used to convert
absorption to BC mass concentration.
Incandescence is the second approach used to
quantify BCa. Laser induced incandescence (LII)
subjects particles in an air stream to a high-intensity
laser in the infrared. Some LII techniques can
measure individual particles, providing data on
particle size, BCa mass concentration (based on
an assumed BC density), and an indication of the
mixing state of the particles. LII is currently used in
limited research applications in the United States.
A1.2.3 Inter-comparisons Among
Optical BCa and Thermal-Optical ECa
Measurements
Given that ECa concentrations are commonly used
to represent BCa, and vice versa, the relationship
between BCa and ECa is important to characterize.
254 Report to Congress on Black Carbon
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Ambient and Emissions Measurement of Black Carbon
Table A1-2. Examples of Commercially Available Optical BCa Measurement Techniques. (Source: U.S. EPA)
Instrument (Manufacturer)3 ^off-l'me (O)™ ^Stream^A? " Wavelengths Measured (See Chapter 2)b
Aethalometer (Magee Scientific)
Particle Soot Absorption Photometer
(Radiance Research)
Multi-Angle Absorption Photometer
(Thermo Scientific)
Transmissometer (Magee Scientific)
Densitometer (Tobias Associates Inc.)
Smoke Stain Reflectometer (Diffusion
Systems, Ltd.)
Hybrid Integrating Plate/Sphere
Photoacoustic soot spectrometer
(Droplet Measurement Technologies,
Desert Research Institute)
Single particle soot photometer
(Artium Technologies, Droplet
Measurement Technologies)
Semi-continuous Field Analyzer (Sunset
Labs)
Photoacoustic Micro Soot Sensor (AVL)
R
R
R
0
0
0
0
R
R
R
R
F
F
F
F
F
F
F
A
A
F
A
370 nm, 880 nm standard
370, 470, 520, 590, 660, 880 and 950 optional
467, 530 and 660 nm
670 nm
370 nm and 880 nm
400 - 650 nm; peak at 575 nm
Monochromatic light; wavelength not specified
633 nm
405,532,781 nm
1064nm
632 nm
808 nm
a The use of commercial trade names or vendor names does not constitute an endorsement by the U.S. EPA.
b A variety of mass absorption coefficients (MACs) or similar conversion factors are used to convert light absorption at a particular wavelength
to BC mass concentration. See the BCa:ECa comparison in Table Al-3 for the MACs used in the comparison studies referenced.
It should be noted that the two measurements are
not always entirely independent, as the selected
conversion factor to estimate BCa is sometimes
based on experiments establishing a relationship
between light absorption and ECa. A number of
inter-comparison studies have examined several
different BCa or ECa measurement approaches
simultaneously to evaluate how well they agreed
(Table Al-3). Recent studies, published in the year
2000 or later, that compare ambient BCa and ECa
measurements were reviewed (Chow et al., 2009;
Bae et al., 2007; Hitzenberger et al., 2006; Snyder
and Schauer, 2007; Sharma et al., 2002; Sahu et al.,
2009; Yang et al., 2006; Miyazaki et al., 2008; Babich
et al., 2000; Ram et al., 2010; Husain et al., 2007;
Jeong et al., 2004; Lim et al., 2003; Hagler et al.,
2007a). In a wide variety of environments, ranging
from the remote Arctic to urban cities, BCa and ECa
measurements were reported to have consistently
high correlation (average R = 0.86 +/- 0.11). In
addition, Figure Al-2 shows that ratios of BCa/ECa
are typically near 1 (BCa/ECa = 0.7-1.3, or within 30%,
for 70% of studies), however there do exist studies
reporting very low BCa/ECa ratios (~0.5) and very
high BCa/ECa ratios (~2).
The ratio of BCa to ECa and the consistency of the
relationship may depend on the aerosol mixture
and/or the specific method used. The difference
in BCa and ECa concentration may also be largely
influenced by the conversion factors used to change
light absorption into mass concentrations for optical
methods as well as corrections for measurement
artifacts. The differences between BCa and ECa
may also be due to a lack of consistency in the
post-processing of the raw measurements among
studies (Venkatachari et al., 2006; Collaud Coen
et al., 2010; Park et al., 2010; Virkkula et al., 2007;
Chow et al., 2009; Bond et al., 1999; Chen et al.,
2004; Jimenez et al., 2007). It should be noted that
these inter-comparison data are based on ambient
measurements and similar data are needed for
source measurements.
Report to Congress on Black Carbon
255
-------
Appendix 7
Table A1-3. Inter-comparison of Ambient BCa and ECa Measurements. Comparisons include (a) BCa
measurements using different instruments; (b) BCa measurements using different instruments after
application of correction algorithm; (c) ECa measurements using different instruments; and (d) BCa
measurements compared to ECa measurements. (Source: U.S. EPA)
Table Al-3 (a) BCa-BCa comparison.
Citation Instrument A Instrument B r r2 ... .,. , Notes
Chow etal. (2009)
Chow etal. (2009)
Chow etal. (2009)
Chow etal. (2009)
Chow etal. (2009)
Chow etal. (2009)
Snyderand Schauer
(2007)
7-AE (660 nm)
7-AE (660 nm)
PSAP (660 nm)
7-AE (520 nm)
PSAP (530 nm)
MAAP (670 nm)
Aethalometer
PSAP (660 nm)
MAAP (670 nm)
MAAP (670 nm)
PA(532nm)
PA(532nm)
PA(670nm)
PSAP
0.98
0.99
0.99
0.96
0.95
0.98
0.93
0.86
1.28
3.52
2.68
4.68
3.69
1.51
1.41
2.68
Fresno Supersite, CA
Fresno Supersite, CA
Fresno Supersite, CA
Fresno Supersite, CA
Fresno Supersite, CA
Fresno Supersite, CA
Slope of line (intercept small)
Overall Average Ratio
Table Al-3 (b) BCa-BCa comparison for study data with correction algorithms applied.
Citation Instrument A Instrument B r r2 ... .,. , Notes
Chow etal. (2009)
Chow etal. (2009)
Chow etal. (2009)
Chow etal. (2009)
Chow etal. (2009)
7-AE adj (660 nm)
7-AE adj (660 nm)
PSAP adj (660 nm)
7-AE adj (660 nm)
PSAP adj (530 nm)
PSAP adj (660 nm)
MAAP (670 nm)
MAAP (670 nm)
PA(532nm)
PA(532nm)
0.95
0.97
0.97
0.95
0.95
1.02
0.9
0.81
1.24
1.17
1.03
Fresno Supersite, CA
Fresno Supersite, CA
Fresno Supersite, CA
Fresno Supersite, CA
Fresno Supersite, CA
Overall Average Ratio
Table Al-3 (c) ECa-ECa comparisons.
c. . High Low Ratio ...
utation Measurement Measurement r n (High/Low) Notes
Bae etal. (2007)
Bae etal. (2009)
Cheng etal.
(2010)
Cheng etal.
(2010)
Cheng etal. (2011)
Cheng etal. (2011)
Cheng etal. (2011)
Cheng etal. (2011)
Chow etal. (2006)
NIER-ECNIOSHTOT
IMPROVE TOR
IMPROVE_ATOR
IMPROVE_ATOR
IMPROVE TOR
IMPROVE TOR
IMPROVE TOR
IMPROVE TOR
IMPROVE_ATOR
UT-EC NIOSH TOT
ACE-Asia TOT
IMPROVE_ATOT
IMPROVE_ATOT
NIOSH TOT
NIOSH TOT
NIOSH TOT
NIOSH TOT
STN NIOSH TOT
0.99
0.79
0.95
0.95
0.97
0.92
0.85
0.81
223
709
89
89
81
97
75
80
17
1.11
2.14
1.74
1.83
1.72
2.78
3.57
4.00
1.87
Semicontinuous Sunset with
different temperature protocols:
NIER - shortened protocols, UT:
nine-step
St. Louis Supersite, MO
TOT and TOR from single DRI Model
2001 analyzer with denuder
TOT and TOR from single DRI Model
2001 analyzer without denuder
Birmingham, AL urban SEARCH site
Jefferson Street, Atlanta, GA urban
SEARCH site
Pensacola, FL coastal SEARCH site
Centerville, AL rural SEARCH site
Fresno Hi-Vol Summer
256
Report to Congress on Black Carbon
-------
Ambient and Emissions Measurement of Black Carbon
Citation M High t M Low t r n m.R*i° . Notes
Measurement Measurement (High/Low)
Chow etal. (2006)
Chow etal. (2006)
Chow etal. (2006)
Chow etal. (2006)
Chow etal. (2006)
Chow etal. (2009)
Chow etal. (2009)
Chow etal. (2009)
Chow etal. (2009)
Chow etal. (2009)
Fujita etal. (2007)
Fujita etal. (2007)
Can etal. (2010)
Klouda etal.
(2005)
IMPROVE_ATOR
IMPROVE_ATOR
IMPROVE_ATOR
IMPROVE_ATOR
IMPROVE_ATOR
IMPROVE_A_TOR_EC
STN_TOR EC
IMPROVE_ATOREC
IMPROVE_ATOREC
IMPROVE_ATOREC
IMPROVE TOR
IMPROVE TOR
IMPROVE TOR
IMPROVE TOR
STNNIOSHTOT
STNNIOSHTOT
STNNIOSHTOT
STNNIOSHTOT
STNNIOSHTOT
IMPROVE_A_TOT EC
STN_TOT EC
STNTOR EC
French two step EC
Sunset Field EC TOT
NIOSHTOT
NIOSHTOT
NIOSHTOT
STN-NIOSHTOT
0.95
0.9
0.94
0.9
0.87
0.94
0.99
8
17
5
17
5
49
18
18
8
48
14
21
6
99
2.86
1.49
2.17
1.46
1.69
1.30
1.41
1.10
1.03
1.82
1.60
1.20
1.09
1.66
1.89
Fresno Hi-Vol Winter
Fresno RAAS Summer
Fresno RAAS Winter
Fresno RAAS Summer with denuder
Fresno RAAS Winter with denuder
Fresno Supersite, CA
Fresno Supersite, CA
Fresno Supersite, CA
Fresno Supersite, CA
Fresno Supersite, CA
Ambient (urban)
Ambient (on-road)
Submarine diesel PM
RM 8785 suspended PM
Overall Average Ratio
Table Al-3 (d) is on the following two pages.
Report to Congress on Black Carbon
257
-------
l/l
00
Table Al-3 (d) BCa-ECa comparisons.
Citation BCa Method (m2^1) l/nm EC Method r Avg BCa Avg EC g^g/EC Location
Chow etal. (2009)
Chow etal. (2009)
Chow etal. (2009)
Bae etal. (2007)
Bae etal. (2007)
Bae etal. (2007)
Bae etal. (2007)
Jeong etal. (2004)
Jeong etal. (2004)
Jeong etal. (2004)
Jeong etal. (2004)
Hagler etal. (2007)
Hitzenberger(2006)
Hitzenberger(2006)
Hitzenberger(2006)
Hitzenberger(2006)
Hitzenberger(2006)
Hitzenberger(2006)
Hitzenberger(2006)
Hitzenberger(2006)
Hitzenberger(2006)
Hitzenberger(2006)
Hitzenberger(2006)
Hitzenberger(2006)
Snyder and Schauer (2007)
Snyder and Schauer (2007)
Sharma etal. (2002)
Sharma etal. (2002)
Sharma etal. (2002)
AethalometerAE-31 PM2.5
MAAP PM2.5
Sunset Optical BC PM2.5
Aethalometer AE-16 PM2.5
Aethalometer AE-16 PM2.5
Aethalometer AE-16 PM2.5
Aethalometer AE-16 PM2.5
Aethalometer AE-20
Aethalometer AE-20
Sunset Optical BC PM2.5
Sunset Optical BC PM2.5
PSAP
AE-9
MAAP
Integrating sphere
Light transmission-white
light
AE-9
MAAP
Integrating sphere
Light transmission-white
light
Aethalometer AE-9
MAAP
Integrating sphere
Light transmission-white
light
PSAP
AethalometerAE-31
Aethalometer AE-11 PM2.5
PSAP PM2.5
Aethalometer AE-11 PM2.5
16.6
6.6
16.6
16.6
16.6
16.6
16.6
16.6
16.6
16.6
b
19
6.5
Calcc
19
6.5
Calcc
19
6.5
Calcc
b
b
19
10
19
660
670
660
880
880
880
880
880
880
660
660
565
830
670
550
white
830
670
550
white
830
670
550
white
565
880
880
565
880
IMPROVE_A_TOR PM2.5
IMPROVE_A_TOR PM2.5
IMPROVE_A_TOR PM2.5
Sunset PM2.5 hourly
Sunset PM2.5 hourly
Sunset PM2.5 hourly
Sunset PM2.5 hourly
Sunset PM2.5 every two hrs.
Sunset PM2.5 every two hrs.
Sunset PM2.5 every two hrs.
Sunset PM2.5 every two hrs.
NIOSH TOT
Cachier two step 1000Cin02
Cachiertwo step lOOOCin 02
Cachier two step lOOOCin 02
Cachiertwo step lOOOCin 02
VDI 650Cin02
VDI 650Cin02
VDI 650Cin02
VDI 650Cin02
TOT800Cin02
TOT800Cin02
TOT800Cin02
TOT800Cin02
Sunset PM2.5 hourly
Sunset PM2.5 hourly
IMPROVE TOR PM2.5
IMPROVE TOR PM2.5
IMPROVE TOR PM2.5
0.89
0.96
0.87
0.93
0.92
0.80
0.70
0.92
0.77
0.97
0.85
0.95
0.72
0.91
0.86
0.89
0.66
0.88
0.78
0.79
0.61
0.88
0.67
0.83
0.91
0.93
0.89
0.99
0.98
0.94
0.95
0.52
0.59
0.59
1.89
1.89
0.9a
0.9
0.3a
0.4a
1.01
0.95
1.01
0.68
0.74
2.18
2.3
0.4a
0.4a
0.4a
0.4a
7ngm 3
0.58a
0.58a
1.42a
0.93
1.00
0.51
0.87
0.80
0.87
0.82
2.25
2.25
0.75
1.00
b
1.14
1.20
0.98
1.20
0.95
1.05
0.84
1.05
1.11
1.11
0.93
1.13
b
b
Fresno, CA
Fresno, CA
Fresno, CA
Gosan, Korea
Gosan, Korea
Gosan, Korea
Gosan, Korea
Rochester, NY
Philadelphia, PA
Rochester, NY
Philadelphia, PA
Greenland - no BC
mass
Vienna, Austria
Vienna, Austria
Vienna, Austria
Vienna, Austria
Vienna, Austria
Vienna, Austria
Vienna, Austria
Vienna, Austria
Vienna, Austria
Vienna, Austria
Vienna, Austria
Vienna, Austria
Riverside, CA
Riverside, CA
Egbert, Canada
Egbert, Canada
Downsview, Canada
3
a.
5'
-------
Citation BCa Method (m2^1) l/nm EC Method r Avg BCa Avg EC g^g/EC Location
Sharmaetal. (2002)
Sharmaetal. (2002)
Sharmaetal. (2002)
Sharmaetal. (2002)
Sharmaetal. (2002)
Sharmaetal. (2002)
Sharmaetal. (2002)
Sharmaetal. (2002)
Sharmaetal. (2002)
Venkatacharietal. (2006)
Venkatacharietal. (2006)
Venkatacharietal. (2006)
Sahuetal.(2009)
Yang etal. (2006)
Miyazakietal. (2008)
Babich etal. (2000)
Babich etal. (2000)
Babich etal. (2000)
Babich etal. (2000)
Babich etal. (2000)
Babich etal. (2000)
Ram etal. (2010)
Husain etal. (2007)
Lim etal. (2003)
Lim etal. (2003)
Lim etal. (2003)
Lim etal. (2003)
PSAP PM2.5
Aethalometer AE-11
PSAP
Aethalometer AE-11 PM2.5
PSAP PM2.5
Aethalometer AE-11 PM2.5
PSAP PM2.5
Aethalometer AE-11 PM2.5
PSAP PM2.5
Aethalometer AE-20 PM2.5
Aethalometer AE-20 PM2.5
Aethalometer AE-20 PM2.5
PSAP PM2.5
Aethalometer AE-16 PM2.5
COSMOS
Aethalometer AE-20
Aethalometer AE-20
Aethalometer AE-20
Aethalometer AE-20
Aethalometer AE-20
Aethalometer AE-20
Aethalometer
Aethalometer AE-21 PM3.2
PSAP PM2.5
PSAP PM2.5
Aethalometer AE-16 PM2.5
Aethalometer AE-16 PM2.5
10
19
10
19
10
19
10
19
10
16.6
16.6
16.6
8.9
16.6
9.8
19
19
19
19
19
19
16.6
16.6
10
10
12.6
12.6
565
880
565
880
565
880
565
880
565
880
880
880
565
880
565
880
880
880
880
880
880
880
880
565
565
880
880
IMPROVE TOR PM2.5
Cachier two step EC 1 1 0OC
in 02
Cachier two step EC 1 1 0OC
in02
NIOSHTOTPM2.5
NIOSHTOTPM2.5
NIOSHTOTPM2.5
NIOSHTOTPM2.5
NIOSHTOTPM2.5
NIOSHTOTPM2.5
Sunset PM2.5 hourly
R&P 5400 PM2.5 hourly
CSN TOT PM2.5
Sunset PM2.5 hourly
IMPROVE TOR PM2.5
Sunset PM10 hourly
IMPROVE TOR PM2.5
IMPROVE TOR PM2.5
IMPROVE TOR PM2.5
IMPROVE TOR PM2.5
IMPROVE TOR PM2.5
IMPROVE TOR PM2.5
Sunset TOT NIOSH PM10
Sunset TOT NIOSH PM2.5
R&P 5400 PM2.5 hourly
RU/OGI TOT PM2.5 hourly
Sunset predecessor
R&P 5400 PM2.5 hourly
RU/OGI TOT PM2.5 hourly
Sunset predecessor
0.69
0.91
0.93
0.92
0.96
0.89
0.89
0.92
0.54
n/a
n/a
n/a
0.98
0.72
0.96
0.87
0.98
0.95
0.95
0.96
0.92
0.79
0.84
n/a
n/a
n/a
n/a
1.01
1.01
1.01
1.18
16.5
n/a
1.1
1.2
0.8
1.1
3.1
1.6
4.45
n/a
1.26
1.26
2.61
2.61
1.42a
0.087,
0.01 2a
0.087,
0.01 2a
1.95a
1.95a
1.82a
1.82a
1.48a
1.48a
0.85
0.55
0.53
n/a
12
n/a
1.4
1.5
1.3
1.5
3.9
1.9
3.84
n/a
2.8
2.33
2.8
2.33
1.2
1.8
1.9
1.0
1.4
n/a
0.79
0.80
0.62
0.73
0.79
0.84
1.2
1.3
0.5
0.5
0.9
1.1
Downsview, Canada
Alert, Canada
(Arctic)
Alert, Canada
(Arctic)
Evans Ave, Canada
Evans Ave, Canada
Palmerston, Canada
Palmerston, Canada
Winchester, Canada
Winchester, Canada
New York City, NY
New York City, NY
New York City, NY
Jeju Island, South
Korea
Xi'an, China
Thailand
Bakersfield.CA
Chicago, IL
Dallas, TX
Philadelphia, PA
Phoenix, AZ
Riverside.CA
Kanpur, India
Lohore, Pakistan
Atlanta, GA
Atlanta, GA
Atlanta, GA
Atlanta, GA
KJ
l/l
vo
a Median concentration.
b BC data presented as absorption coefficients (Mm *). Ratio of BCa/ECa and linear regression equations not extracted for these papers, although it could be calculated.
c Calibration curve based on dissolved carbon black.
C-
O
3
-------
Appendix 7
14
12
10
30%
01
01
2 -
o
CO
0
o
LT>
0%
O O O
-------
Ambient and Emissions Measurement of Black Carbon
in Watson et al. (2005) showed differences of a
factor of 2 to be common for measurements of
ECa. Method differences have been found not
only to depend on the operational definitions of
the methods, but also to depend on the chemical
composition and source of aerosol collected (i.e.,
location and seasons), particle loading on the
filter, and uniformity of the filter deposit. Samples
containing biomass smoke present more difficulties
due to components (inorganic compounds) that
result in an overestimation of ECa (Novakov and
Corrigan, 1995a). The presence of BrC can also affect
the measurement of ECa. Reisinger et al. (2008)
evaluated the BrC content of samples and the impact
on the comparison of thermal-optical methods
and found TOT methods to be less sensitive to the
presence of BrC than other thermal methods tested.
There are currently no reference materials that reflect
the variety of aerosol types in the atmosphere and
there is no standardized method protocol. There is
one consensus-based standard reference material
(SRM 8785) available from the National Institute of
Standards and Technology (NIST) that has values
assigned for OCa and ECa as measured by the two
EPA protocols discussed below (IMPROVE TOR
and NIOSH-like TOT). The ECa assigned by the two
protocols (IMPROVE: NIOSH) in the NIST SRM differ
by a factor of 1.7 (Klouda et al., 2005).
Results from thermal-optical ECa:ECa comparison
studies that use methods similar to the two EPA
protocols that are not summarized in Watson et al.,
(2005) and published since 2005 were reviewed and
summarized in Table Al-3 (Klouda et al., 2005; Bae et
al., 2007; 2009; Cheng et al., 2010; 2011; Chow et al.,
2006; 2009; Fujita et al., 2007; Can et al., 2010). The
average ratio of ECa:ECa from these studies is 1.9.
Chow et al. (2006) also compared ECa measurements
by IMPROVE and NIOSH-like TOT methods for
a variety of lab generated source samples. The
IMPROVE:NIOSH-like TOT ECa range of ratios were
from 1.01 to 1.04 for diesel, acetylene flame, carbon
black, and graphite source samples, and 1.13 for
electric arc samples. The biggest difference was
found for the wood smoke source sample (ratio of
1.88).
A1.2.5 Inter-comparison of Two EPA ECa
Measurement Protocols
The IMPROVE TOR and NIOSH-like TOT methods
have been widely used in the EPA's national urban
Chemical Speciation (CSN) and rural IMPROVE
ambient monitoring networks. EPA has transitioned
the urban CSN from the NIOSH-type TOT method to
the IMPROVE_A TOR method. The transition began
in May 2007 and was completed in October 2009
and includes a change to the sampling system as
well as the analytical method. The major difference
in the sampling method is the sampling flow rate
(increased to ~22 LPM from ~6.7 LPM) and sample
filter diameter (reduced from 46.2 mm to 25 mm),
which results in an overall increase in pressure drop
across the filter during sampling. The combination
of these changes results in a reduction in the OC
measured, which is most likely related to a change
in sampling artifacts. The rationale for the transition
of the urban CSN to IMPROVE-like sampling and
analysis method was to institute consistency in the
carbon measurements across the EPA's national
particulate monitoring networks.
To understand the differences between the two
carbon monitoring protocols, EPA established
pairs of old and new CSN monitors at 11
sampling locations and collected parallel carbon
measurements for 12 months from May 2009 to
April 1010. Most other CSN sites also collected 2
months of parallel measurements when they initially
transitioned between May 2007 and October 2009.
In addition, lower flow rate CSN samplers whose
carbon was measured with the NIOSH-type TOT
method were used to collect data at 14 urban sites
in the IMPROVE network.
The comparison between the previous CSN TOT
data and the current CSN IMPROVE TOR data
indicates that measured EC is reasonably consistent
between the methods at the 11 locations that
produced 2009-2010 data (Figure Al-3). These
data suggest that both monitoring protocols could
be interchangeably used to evaluate BC aerosols
predicted by climate and air quality models, and to
evaluate trends. The seasonal differences in these
EC differences are modest, and may be related to
the combined effect of sampling rates and analytical
protocol and the resulting differences in measured
OC as described above. However, when all the
parallel EC measurements are considered, a different
pattern emerges.
The comparison of urban EC derived with the
IMPROVE TOR method to that from the CSN NIOSH-
like TOT method in Figure Al-4 shows the ratio of
monthly values decreasing from approximately 1.5
in 2005 to approximately 1.0 towards the end of
2009. The higher monthly ratios observed between
2005-2006 are consistent with the finding that
CSN EC is on average 30% lower than IMPROVE EC
(Hand et al., 2011) and the ratios shown for the last
12 months are consistent with the data presented
Report to Congress on Black Carbon 261
-------
Appendix 7
1
csL
O
4-
3-
2-
1 -
0-
1.0
-0.5
-0.0
--0.5
U
LU
0)
CT
ra
I
3 4
7 8 9 10 11 12
month
Figure A1-3. Monthly Distribution of ECa/ECa Ratio For Two EPA methods
(TOR/TOT) From 897 Collocated Measurements Among 11 Urban CSN
Locations. Average CSN NIOSH ECa concentration is red; IMPROVE TOR
ECa is blue; the distribution of daily ratios is presented as box plots (black).
(Source: U.S. EPA, AQS)
in Figure Al-3. Although the number of included
monitors varies from month to month, the pattern
from a consistent set of 5 collocated IMPROVE-
CSN sites from 2005 thru 2009 reveals the same
general trend. Although the ratio appears to have
decreased during this five year period, the cyclical
behavior suggests that relatively higher ratios often
occur during the warmer months. EPA will continue
to evaluate the differences between the two
measurement protocols and possible connections
to changes in the way the measurements were
conducted as well as the potential influence of
changes in other collected aerosols.
A second implication of the change from the
NIOSH-like TOT to the IMPROVE TOR monitoring
method along with the change in samplers, relates
to measured OCa and its sampling artifacts. In some
cases, sample collection procedures can lead to the
inclusion of positive artifacts—mistakenly measuring
non-PM components such as vapors as if they
were in fact carbonaceous PM. Other procedures
can lead to the exclusion of relevant material,
producing negative artifacts. These artifacts are a
problem particularly for measuring concentrations
of OCa; sampling artifacts for EC are thought to be
negligible, simply because the EC collected on the
filter is more stable (non-reactive or volatile).
"1-5 Because sampling artifacts
are most likely to affect
measurements of OCa, they
may be most important for
understanding OCa/ECa
ratios (i.e., representing OC/
BC). Figure Al-5 shows the
monthly distribution of OC/
BC among 897 measurements
at 11 urban monitoring sites2
that concurrently sampled with
two alternative measurement
protocols (NIOSH TOT and
IMPROVE TOR) during 2009-
2010. Though ECa can vary
somewhat according to the
monitoring protocol (see
further discussion of NIOSH
TOT and IMPROVE TOR), OCa
can vary even more widely as
a result of the correction used
for OCa sampling artifacts. As
the figure shows, the OCa/ECa
ratios with the CSN NIOSH TOT
method have large seasonal
variation and for the 11-site
group, the median value is
as high as 5. On the other hand, the CSN TOR OC/
EC ratios do not display strong seasonality and
have monthly median values of ~2-2.5. The latter
are more consistent with average estimated direct
emission OCa/ECa levels described in Chapter 5, as
well as with the artifact corrected ratios described
elsewhere (Novakov et al., 2005). However, they
do not display the seasonal change in OCa/ECa
ratios due to secondary organic aerosol (SOA)
reported elsewhere. As discussed in Chapter 2
and Chapter 5, the correct characterization of OCa
aerosol is critical for differentiating among reflecting
vs. absorbing particles for assessment of radiative
forcing, where OC is assumed to be mainly light
scattering. While the IMPROVE TOR OCa is adjusted
for sampling artifact with backup filters, the CSN
NIOSH TOT protocol is only adjusted with nominal
network value of 1 ug/m3 which may be too low
(Chow et al., 2010a). On the other hand the much
more suppressed seasonal behavior in the TOR
carbon ratios could be related to the higher flow
rate IMPROVE protocol samplers which may not
fully retain semi-volatile OC particles. The latter will
require further study to understand its implications
for using these measurements to develop emissions
211 site inter-comparison group includes Bronx and Queens, NY;
Atlanta GA, Birmingham AL, Detroit MI, Cleveland OH; Chicago IL,
Denver CO, LA (Rubidoux), CA; Sacramento CA and Seattle, WA.
262
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Ambient and Emissions Measurement of Black Carbon
CO
9 1.75
g 1.50
H,
(3 1-25
LLJ
f 1.00
a
S> 0.75
§ 0.50
0.25
+ 1.75
JAN05 JAN06 JAN07 JAN08 JAN09 JAN10 JAN11
Figure A1-4. Trend in Ratio of Urban EC Derived From IMPROVE TOR and CSN NIOSH-like TOT Methods. The ratio
of monthly averages derived from collocated IMPROVE (or CSN IMPROVE-like) with CSN MetOne samplers is shown,
where the former measurements are produced with the IMPROVE protocol and the latter measurements are produced
with the NIOSH-like TOT protocol. The results provided by ECa from the IMPROVE network samplers are shown with (*)
and those from the CSN are represented by squares. Months October-March are denoted in blue and months April-
September are shown in red. The dotted line is a spline fit through the monthly data. (Source: U.S. EPA, AQS)
g
ro
U
LU
U
O
10-
9-
8-
7-
6-
5-
4-
3-
2-
1 -
0-
05
O-
u
LU
U
o
10-
9-
8-
7-
6-
5-
4-
3-
2-
1 -
0-
(b)
8 9 10 11 12
1234567
9 10 11 12
Figure A1-5. Monthly Distribution of OC/BC ratios for 11 CSN Sites Produced With (a) the NIOSH-like
TOT and (b) IMPROVE_A TOR Monitoring Protocols, 2009-10. Nominal OC sampling corrections of 1 ug/
m3 for CSN NIOSH type TOT have been used (Frank, 2006). The IMPROVE protocol data are adjusted with
backup filters. Due in part to inability to adequately correct the CSN NIOSH OC sampling artifacts (Chow et
al., 2010a), these data may in fact overstate ambient OC/BC and imply a seasonal pattern which may be an
artifact of the monitoring method. (Source: U.S. EPA)
Report to Congress on Black Carbon
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Appendix 7
Distribution of Monthly Average EC
as measured in CSN at 15 sites
NIOSHEC
o
o
HI
Figure A1-6. Monthly Distribution of Reported CSN EC at
15 sites From 2002 Through 2010. The ECa data produced
with the CSN NIOSH-like TOT is shown in red and the ECa data
produced with the IMPROVE_ATOR monitoring protocol is
shown in blue. (Source: U.S. EPA, AQS)
CO
E
o
ro
I
o
o
O
O
CO
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
Figure A1-7. Ambient BC Trends (2002-2010), Based on
Monthly Distribution of Average ECa Concentrations Among
15 CSN Monitoring Locations in the United States. The map
shows the location of the 15 monitoring sites. (This figure is also
as shown as Figure 5-12.) (Source: U.S. EPA)
inventories and to evaluate climate
modeling data.
A1.2.6 Implications of
Changes in ECa Monitoring
on Estimated Concentration
Trends
The urban ECa measurements as
reported by the CSN in Figure
Al-6 reveal a different picture
than the one presented in Chapter
5, Figure 5-12. (For convenience,
Figure 5-12 is also re-produced
as Figure Al-7.) The reported EC
data based on the NIOSH-type
measurements in Figure Al-6
appear to depict a slight upward
progression from 2002 thru 2006.
However, when the data are
adjusted using a 5-month moving
average of the monthly ratios for
2005-2010 shown in Figure Al-4,
together with the 2005 average
ratio for earlier data, an estimated
downward trend in EC is revealed.
A1.2.7 Other Measurements
Microscopy (the use of
microscopes to view the structure
of particles) and spectroscopy
(measurement of a chemical
as a function of wavelength)
provide additional information
about the physical and chemical
structure of carbonaceous PM.
An advantage of these methods
is that they provide detailed
information about how particles
age and transform from the point
of emission to the atmosphere. A
variety of microscopy techniques
have been applied to investigate
carbon particles. Scanning
electron microscopy (SEM)
with energy dispersive X-ray
spectroscopy (EDX), transmission
electron microscopy (TEM),
and Raman Microspectroscopy
(RM) are the most widely used
and have provided the most
significant information about
carbon aerosol composition to
date. Like the thermal-optical
264
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Ambient and Emissions Measurement of Black Carbon
and light absorption measurement methods,
microscopy has limitations and is subject to artifacts
and interpretation issues, but these techniques do
provide additional information not gathered by the
thermal and optical measurement techniques.
A1.2.8 Limitations of Ambient Measurement
Methods
Specific operating conditions, such as the heating
temperature, time of heating, and char correction
procedures, can influence thermal-optical
measurement results. Chemical composition and
emissions sources of the measured aerosol, filter
loading, and uniformity of the filter particle deposit
can also influence OCa and ECa values obtained.
Studies suggest that EC measurements for some
types of emissions (biomass smoke, dust) may be
more strongly affected than traffic-related (e.g.,
diesel) samples, in part because of higher levels
of inorganic components and BrC (Novakov and
Corrigan, 1995b). A summary of the comparison
of optical BCa to thermal ECa measurements is
provided above.
Currently, there are no reference standards for
assessing the accuracy of OCa or ECa measurements
by thermal methods, nor is there a standardized
method protocol for distinguishing between OCa
and ECa. Development of standard reference
materials and the consensus on standardized
method protocols (including data reporting
procedures) will be important in the future for the
consistent measurement of OCa and ECa for climate
purposes.
All optical BC measurements share a fundamental
limitation in that they do not directly measure BC
mass concentration. Instead, conversion factors
(e.g., mass absorption efficiency or mass absorption
cross-section) are necessary to generate BCa
mass concentrations from the different optical
measurements. In addition, the most commonly
used filter-based methods are prone to artifacts
during sampling. The extent of filter loading can
influence particle scattering and shadowing effects
which bias results (Park et al., 2010; Bond et al.,
1999; Weingartner et al., 2003). While several filter-
loading-based correction algorithms have been
introduced (e.g., Virkkula et al., 2007), it is uncertain
as to whether a correction algorithm should be
universally applied as the artifacts may depend upon
the particle composition and concentration. Because
the aerosol absorption and derived BCa depend on
wavelength, it should be noted that some reported
BC that is based on wavelengths in the visible
spectrum may include other LAC.
A1.2.9 Critical Gaps and Research Needs in
Ambient Measurement Methods
In light of the limitations discussed above, the
following research can help improve the ambient
measurement of BC and LAC in the future and
reduce the uncertainty:
• Further comparisons of the predominant thermal
and optical methods in use today are needed
to better understand and characterize the
differences and uncertainties. As comparisons
are made, it is important to clearly document the
operational conditions of the methods used.
• Having a consistent, well-defined "BC" reference
material would help to better understand method
differences and define the uncertainties in the
various measurement methods.
• It is important to agree on a standardized
method of operation and calibration for those
methods identified as most important for
measuring BC in support of climate and health.
• There is a need to develop methods capable of
quantifying particulate components, referred
to as BrC or (collectively with BC) as LAC, that
provide additional light absorption in the near-
UV and UV wavelengths.
• To ensure proper use of measurements,
consistent data reporting (including metadata)
of the sampling and analysis protocols and data
adjustments must be provided.
• Continued research and further development
is needed for continuous or real-time single
particle measurements (e.g., aerosol time-of-
flight mass spectrometry and single particle
soot photometers) to enhance our knowledge of
particle composition and mixing state.
A1.3 Black Carbon Emissions Source
Measurements
Source measurements are used for a variety
of purposes, including regulatory compliance.
However, in the United States and elsewhere, such
Measurements generally focus on total PM2.5 mass.
measurements of specific components are not
required in the United States as part of regulatory
Report to Congress on Black Carbon 265
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Appendix 7
testing, and EPA does not have an official source
measurement method. Instead, PM composition is
measured largely for research purposes, including
development of EPA's emissions models. Available
source measurements are also used to develop
and verify emissions inventories, refine standard
measurement approaches, and assess control
technologies and mitigation approaches. Due to
the limited amount of source emissions data for the
carbonaceous content of PM, EPA often must rely
on data and methodologies for total PM mass, or
substitute emissions models.
A1.3.1 Stationary Source Emissions
Measurement Methods
Most current federal stationary source emission
standards are focused on the regulation of filterable
total PM mass. For most stationary sources in the
current inventory, PM2.5 emissions are derived
from use of a scaling factor applied to collection
of filterable total PM and the PMi0 size fractions.
Some local/state and site specific standards also
require testing for PMi0 and PM2.5 mass, which
sometimes includes both size fractions of filterable
and condensable PM. The latter allows for inclusion
of certain semi-volatile particles. EPA has recently
promulgated a stationary method for PM2.5 mass
and refined the condensable stationary source
Table A1-4. Stationary Source Emissions Measurement Methods. (Source: U.S. EPA)
Method PMType Temperature (°F) Purpose CFR Reference
EPA Method 5
EPA Method 5A
EPA Method SB
EPA Method 5D
EPA Method 5E
EPA Method 5F
EPA Method 5G
EPA Method 5H
EPA Method 51
EPA Method 17
EPA Method 201
EPA Method 201A
EPA Method 202
EPA Conditional Test
Method -039
Filterable
Filterable
Filterable
Filterable
Filterable and Total
Organic Material
Filterable
Filterable and
Condensable
Filterable and
Condensable
Filterable
Filterable
Filterable
10 |jm
Filterable
10 um/2.5 |am
Condensable
Total 10 um/2.5 urn
(Filterable
and Condensable)
248 ± 25
108±18
320 ± 25
248 ± 25
248 ± 25
320 ± 25
<90
<248and <68
248 ± 25
StackTemperature
StackTemperature
StackTemperature
85
85
General
Asphalt Roofing
Utility Plants
Positive Pressure baghouses
Wool Fiberglass
Nonsulfate Filterable PM
Wood Heaters - Dilution
Wood Heaters
Low level general
General
General - Particle Sizing
General - Particle Sizing
General - Condensable PM
General - Dilution based PM
40 CFR 60 Appendix A-3
40 CFR 60 Appendix A-3
40 CFR 60 Appendix A-3
40 CFR 60 Appendix A-3
40 CFR 60 Appendix A-3
40 CFR 60 Appendix A-3
40 CFR 60 Appendix A-3
40 CFR 60 Appendix A-3
40 CFR 60 Appendix A-3
40 CFR 60 Appendix A-6
40 CFR 51 Appendix M
40 CFR 51 Appendix M
40 CFR 51 Appendix M
Example State, VCS, and International Methods
CARB5
CARB501
ASTM D6831 - 05a
ISO 9096 and EN 13284
VDI 2066 Part. 10
method and in the Norm
EN 13284-1
Filterable
Filterable, multiple
aerodynamic sizes
Filterable
Filterable
Filterable
10 um/2.5 urn
248 ± 25
StackTemperature
StackTemperature
General - Particle Size
Continuous PM
http://www.arb.ca.gov/
testmeth/voll/Meth_501.pdf
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Ambient and Emissions Measurement of Black Carbon
measurement protocol (U.S. EPA, 2010f); overtime
this will help ensure greater consistency in stationary
source emissions measurements. However, stationary
source data currently available for PM2.5 inventory
purposes are based on non-standardized methods
and procedures for PMi0 and total filterable PM.
Due to the complex nature and variety of sources,
regulatory and other standardized source PM
methods are mainly designed to provide consistent
results across a certain category of sources and not
necessarily the entire universe of sources (Myers,
2006). Thus, compilations of source emissions
measurements for total PM mass exist such as
EPA's AP-42 (compilation of EPA's emission factors)
and the U.S. National Emissions Inventory (NEI).
However, none of these compilations reflects routine
sampling required by regulation for all sources in
the inventory. Measurement of carbonaceous PM
components including BC or EC are not required as
part of compliance testing. Such results are generally
available only in the academic literature.
There are a large variety of methods for the
measurement of PM mass from stationary sources,
many of which measure both the filterable and
condensable fractions of PM2.5. These methods vary
due to operational differences such as filtration
temperature and conditioning and treatment of the
different components of PM. Table Al-4 provides a
list of commonly used stationary source methods
and some examples of operational differences for
determining PM mass from a variety of sources.
A1.3.2 Mobile Source Emissions
Measurement Methods
Mobile sources consist of a diverse group of vehicles
and engines, including light-duty gasoline vehicles,
heavy-duty diesel trucks, gasoline-powered nonroad
engines (e.g., lawnmowers, snowmobiles, recreational
boats), nonroad diesel engines (e.g., excavators,
locomotives, and marine vessels), and turbine
and propeller-driven aircraft. Due to their diverse
technologies and applications for highway and
nonroad uses, there is considerable variability in BC
emissions from mobile sources.
In the United States, particles in mobile source
exhaust emissions are measured for compliance
with PM emission standards and are expressed on a
mass per unit work (g/bhp-hr) or mass per distance
traveled (g/mi) basis. For regulatory certification,
diesel exhaust particle emissions are measured
using procedures described in 40 CFR Part 1065,
which employs an engine dynamometer paired
with a dilution sampling system collecting samples
on Teflon filters at temperatures of about 125°F
(which reduces water condensation, yet allows for
condensation of organic compounds). The filters
are then conditioned at a specific temperature and
humidity3 and weighed. This procedure is commonly
used to measure PM from non-diesel mobile
sources for research purposes.
Mobile source emissions of BC are almost always
measured as ECa. Unlike PM measurements,
however, ECa measurements are not routinely taken
and EPA does not presently have an official (or
even recommended) EC measurement method for
mobile sources for regulatory purposes. However,
EPA does measure BC in its mobile source emissions
characterization programs. There, BC is measured
as a particulate matter (PM) component for both
gasoline vehicles such as light-duty cars/trucks and
diesel vehicles such as heavy-duty diesel trucks
(up to 80,000 Ibs. gross vehicle weight). It is also
measured to a more limited extent from nonroad
diesel and even gasoline engines (both 2-stroke
cycle engines which have lubricating oil mixed
with the fuel) and 4-stroke cycle engines. It is also
measured in PM from locomotives, commercial
marine, and aircraft.
Sampling temperature has a major effect on the
quantity and even the composition of PM. PM
emissions are collected on a filter from diluted
exhaust. The general methodology for measuring
mobile source PM involves diluting the vehicle
exhaust with ambient air roughly at a 10/1 dilution
ratio (although the dilution ratio varies greatly
depending on engine operating mode) using a
stainless steel dilution tunnel. The filter temperature
is about 125°F, which is high enough to prevent
water condensation on the filter from the copious
amounts of water vapor present in vehicle exhaust
formed from fuel combustion. This temperature
also allows some condensation of the organic
hydrocarbon compounds present in vehicle exhaust.
This general method was developed and has been
in use since about 1970 for both diesel and gasoline
exhaust. This methodology is also used for EPA
emission standards for exhaust from diesel engines
including on-road trucks and more recently nonroad
diesel engines. This measurement system, known
as constant volume sampling of an exhaust stream
that is diluted with varying amounts of dilution air,
allows for accurate mass weighting of emissions
over transient driving conditions (accelerations,
3 Mobile source measurements are made at45%RH, while ambient
measurements and many other source tests use 35%RH.
Report to Congress on Black Carbon 267
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Appendix 7
decelerations, steady-state cruise, and idle) where
exhaust volume varies. In the ambient air though,
vehicle exhaust is rapidly diluted to about 1,000:1
which results in somewhat different condensation of
the hydrocarbon compounds into particulate.
The PM measurement method is more developed for
diesel PM than for gasoline PM. Numerous studies
have been done measuring diesel PM starting with
the first EPA emission standard for the 1970 model
year for visible smoke from diesel engines.
A1.3.3 Use of Emissions Source Test Data
Though carbonaceous components of PM are not
systematically measured across all categories, both
EPA and external researchers have measured these
components from some source categories. EPA
has compiled all available source emissions data
into a database called SPECIATE, which currently
contains 3,326 raw PM profiles. Because many of
these measurements are drawn from research on
emissions measurements, the data comes from
a variety of sampling and analytical technologies
(e.g., see Chang and England, 2004). Despite the
uncertainties and limited size of the testing dataset
compared to the total number of sources, the
SPECIATE database represents the best compiled
source of data available. A subset of these data was
selected to characterize the source profiles for 15
source categories reported in Chapter 4, Figure 4-1.4
The number of individual profiles by source category
can be quite limited and sometimes only a single
value was used. Similar summaries are available
elsewhere (Chow et al., 2010a). Note that for some
sources, the sum of BC and OC is less than 100% of
PM2.5 mass. The raw data used to compile Figure 4-1
is available in Table Al-5, along with the percent of
estimated non-carbon PM and the OC/BC ratios.
As discussed in Chapter 4, however, EPA does not
use any of these profiles for on-road vehicles, since
the mobile MOVES model directly calculates EC
emissions (U.S. EPA, 2010c). Mobile sources have
more variability in emissions than stationary sources,
because mobile-source EC varies with driving mode,
specific model mix, and other conditions. MOVES is
4 Following the procedures of Reff et al. (2009), the raw profiles
in SPECIATE were modified so that all EC was adjusted to be
representative of the TOR analytical method and so that the sum
of the species equals the PM2.5 mass, if the raw profile was not
provided in that format. To provide a more representative median
among available test data, subsets of multiple source tests were first
combined into a composite profile. Some uncertainty in expressing
EC as a fraction of PM2.5 may be related to the water content of PM2.s
mass.
designed to capture this variability. Currently, EPA
still uses speciation profiles for nonroad diesel.
A1.3.4 Limitations of Source Emissions
Measurement Methods
To estimate EC emissions for a specific source
category, EC is typically assumed to be a specific
fraction of PM2.5 and then total PM2.5 mass is used
as the starting point. Thus, the measurement and/or
estimate of PM2.5 mass is one very important source
of potential uncertainty. There are inconsistencies in
the way PM25 is measured among source categories,
including in the approach for determining filterable
and condensable mass, filter equilibration conditions
(including laboratory relative humidity), temperature
of testing, and dilution and related procedures for
semi-volatile PM. Some of these variables can also
affect the measurement of carbon components.
Because of the way estimates of PM components
are generated, both the PM and carbon-specific
measurements can affect estimates of BC and OC
emissions for a given source category.
Current PM25 estimation methods based on PMi0
and total filterable PM can produce variable results,
particularly the methods that include condensable
PM. For certain stationary source categories,
this can produce measurement artifacts that can
overestimate the condensable PM emissions by an
order of magnitude.5
The use of scaling factors applied to filterable total
PM and/or PMi0 to generate estimates of PM25
introduces additional uncertainty to the estimated
emission rate (National Research Council, 2004).
Finally, the representativeness of a particular source
profile based on a limited number of source tests
is questionable, and derived composite profiles
applied to a large number of sources is another
source of uncertainty. For both PM and speciation
test data, there are the related representativeness
issues of tests conducted with actual vs. allowable
emissions from the stacks and effluents; tests
conducted at facilities of varying age and with
different degree and type of controls; and tests
affected by other operating conditions. These
factors are often not taken into account when BC
profiles are applied to PM2.5 emissions. There are
also potential issues regarding PM25 mass closure
(including treatment of volatile components, particle
5 Example artifacts include the potential conversion of sulfur
dioxide gas into sulfate particles, affecting the reported PM mass.
268 Report to Congress on Black Carbon
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Ambient and Emissions Measurement of Black Carbon
bound water) and comparison of BC data based on
different measurement methods.
A1.3.5 Critical Gaps and Research Needs in
BC Emissions Sampling and Measurement
Methods
In light of the limitations discussed above, the
following research can help shed light on amounts
of BC and LAC emitted by various sources and
lessen the uncertainty in developing an inventory of
emissions:
1. For all source measurements
• Understand how the source EC values relate to
source BC values based on currently available
techniques.
• Develop high-quality source profiles for sources
that need improved characterization for BC,
including research into how to quantify the
additional light-absorbing components in the
near-UV or UV spectrum that are referred to as
BrC or, collectively with BC, as LAC.
• Develop a standard "BC" reference material and
establish a standard measurement method to
report source data as BC.
2. Stationary source measurements
• Understand the effect of varying source test
methods and conditions on measured PM2.5 and
BC; and standardization of PM source testing
procedures for filterable and condensable PM.
• Perform uncertainty analysis of all source profiles
that exist in SPECIATE and how the total mass
from the SPECIATE collection methods relates
to the total mass from the methods used in the
emissions inventory.
• Increase the quantity and quality of meta-data
available in the databases that better explain
how PM2.5 and EC fractions were derived for the
various sources in EPA's inventories.
3. Mobile source measurements
• Develop standard measurement methods for BC
for both on-road and nonroad engines, especially
diesels but also gasoline vehicles/engines.
• Establish more routine measurement procedures
for BC, including ones that can measure these
quantities over short time periods (even
instantaneously) as well as over an entire driving
cycle.
Report to Congress on Black Carbon
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SI
o
Table M-5. Data Used to Prepare BC and OC Source Profile Box Plots (Chapter 4, Figure 4-1). OC/BC Ratios and OC+BC as Percent of PM Are Also Included.
(Source: U.S. EPA)
. . Residential
Stats Agricultural Bituminous _ Distillate Oil HDDV LDDV Natural Gas ,. _ Prescribed ProcessGas Wood SubBituminous ,,,.,,,
Charbroilmq . . . Gasoline Gasoline . . . Wildfires Fired
(%ile) Burning Combustion Combustion Exhaust Exhaust Combustion ,_ . r . Burning Combustion Combustion: Combustion „ ..
M Exhaust Exhaust M Boiler
HardSoft
OC Fraction of PM25
10th
25th
50*
75th
90*
N
0.30
0.34
0.40
0.44
0.56
9
0.02
0.02
0.03
0.07
0.12
3
0.34
0.46
0.70
0.84
0.87
4
0.25
0.25
0.25
0.25
0.25
1
0.18
0.18
0.18
0.18
0.18
1
0.20
0.25
0.32
0.39
0.44
4
0.25
0.25
0.25
0.25
0.25
1
0.84
0.84
0.84
0.84
0.84
1
0.30
0.44
0.55
0.67
0.75
10
0.65
0.70
0.71
0.79
0.83
7
0.05
0.20
0.35
0.46
0.57
3
0.39
0.47
0.53
0.58
0.68
12
0.02
0.02
0.03
0.04
0.04
2
0.47
0.47
0.56
0.64
0.64
2
0.33
0.33
0.33
0.33
0.33
1
BC Fraction of PM25
10th
25th
50*
75th
90*
N
0.05
0.08
0.10
0.12
0.13
9
0.01
0.01
0.02
0.05
0.08
3
0.00
0.01
0.02
0.06
0.10
3
0.10
0.10
0.10
0.10
0.10
1
0.77
0.77
0.77
0.77
0.77
1
0.31
0.38
0.53
0.63
0.64
4
0.38
0.38
0.38
0.38
0.38
1
0.01
0.01
0.01
0.01
0.01
1
0.09
0.14
0.19
0.23
0.34
10
0.01
0.01
0.02
0.04
0.07
7
0.10
0.13
0.17
0.22
0.27
3
0.01
0.04
0.06
0.10
0.12
12
0.02
0.02
0.04
0.07
0.07
2
0.03
0.03
0.09
0.16
0.16
2
0.14
0.14
0.14
0.14
0.14
1
OC:BC Ratios
10th
25th
50th
75th
90*
5.9
4.2
4.1
3.6
4.3
1.9
1.7
1.6
1.5
1.4
41.2
31.1
13.5
8.5
2.5
2.5
2.5
2.5
2.5
0.2
0.2
0.2
0.2
0.2
0.6
0.7
0.6
0.6
0.7
0.6
0.6
0.6
0.6
0.6
59.9
59.9
59.9
59.9
59.9
3.3
3.2
2.9
3.0
2.2
54.3
49.4
38.6
19.3
12.0
0.5
1.5
2.1
2.1
2.1
27.6
12.4
9.4
5.9
5.5
1.0
1.0
0.7
0.7
0.7
14.5
14.5
5.9
4.1
4.1
2.4
2.4
2.4
2.4
2.4
BC+OC,as%PM
10th
25th
50th
75th
90*
36%
43%
49%
57%
69%
3%
4%
4%
12%
20%
34%
47%
72%
90%
97%
35%
35%
35%
35%
35%
95%
95%
95%
95%
95%
51%
63%
86%
102%
108%
63%
63%
63%
63%
63%
85%
85%
85%
85%
85%
39%
57%
74%
90%
109%
66%
72%
73%
83%
90%
15%
34%
53%
68%
83%
40%
51%
58%
68%
81%
4%
4%
7%
11%
11%
50%
50%
65%
80%
80%
46%
46%
46%
46%
46%
3
a.
5'
n
o
3
O
3
2
S"
n
>r
n
Q
5-
o
3
-------
Appendix 2
Black Carbon Emissions Inventory
Methods and Comparisons
A2.1 Introduction
This appendix provides specific details on the
approach used to generate domestic inventories
for stationary, area, and mobile sources, and
compares that approach to methods used in
compiling international inventories. It explores
key methodological similarities and differences
between inventories, and also outlines the specific
methodologies and data inputs used to construct
key global and regional inventories currently
available.
In general, existing inventories for BC are based
on calculations rather than actual emissions
measurements. Direct emissions measurements
of BC and other PM constituents are rare, and no
known inventory is based on direct BC emissions
data. Instead, "BC" inventories are calculated
mathematically from PM2.5 inventories. These
calculations divide the direct carbonaceous particle
emissions from the PM2.5 inventory into two
categories: EC and OC. Thus most "BC" inventories
are actually EC inventories. Though sometimes the
terms EC and BC are used interchangeably, EC is
actually more narrowly defined (see Chapter 2 and
Appendix 1). By tracking only EC, current inventories
fail to account for the portion of primary OC
emissions that is light absorbing (including some
BC and also BrC). As discussed in Chapter 4, this
means current domestic and international inventories
systematically underestimate total LAC; however,
the magnitude of this gap has not been adequately
quantified to date.
Most inventories use a bottom-up approach, first
pairing PM2.5 emission factors with activity level data
for the source category to generate PM2.5 emissions
estimates, and then applying a speciation factor to
estimate the amount of BC (or other constituents)
contained in the total mass of PM2 5 emissions.
The BC emissions for individual source categories
are then aggregated to form total BC emissions
estimates. The speciation factors for an individual
source category relate emissions of specific
constituents to total PM2.5 mass—for example, the
amount of BC to total PM25. PM25 emission factors
and the speciation factors for particular constituents
can be based on either fuel consumption data
(i.e., estimated emissions of total PM2.5 or specific
constituents per unit of fuel) or actual measured
source emissions from emissions testing (see
Appendix 1). Some inventories use a combination
of these different approaches, depending on the
information available for each source category.
In a few cases, emissions may be back-calculated
from remote sensing of primary PM25 emissions
or measured ambient data of the amount of
carbonaceous aerosols in the atmosphere. This
kind of top-down approach is far less common;
currently only a few regional inventories in Asia rely
on such methods.
Additional information on approaches used to
calculate the U.S. emissions inventory and other
global/regional inventories is provided in the
following text.
A2.2 Development of U.S. National
Emissions Inventory for Black and
Organic Carbon
EPA's National Emissions Inventory (NEI) is a
bottom-up compilation of estimates of air pollutants
discharged on an annual basis by source category
(see http://www.epa.gov/ttn/chief/eiinformation.html}.
EPA's 2002 Consolidated Emissions Reporting Rule
[http://www.epa.gov/ttn/chief/cerr/cerr.pdf] provides
a regulatory basis for the collection of emissions
inventory information. Currently, emissions of BC
and other PM constituents (OC, nitrates, sulfates,
crustal material) are not directly reported as part
of the NEI. Rather, BC emissions for most sources
are estimated by matching PM2 5 emissions from
the NEI for those sources to source-specific BC
speciation profiles from the "SPECIATE" database
(see Appendix 1 for details on this database). The
one exception is on-road mobile sources, for which
BC emissions are estimated directly through models.
The following sections provide more information on
the specific methods used to compile the inventory
for both stationary and mobile sources. More detail
is provided for mobile source inventories due to the
Report to Congress on Black Carbon
271
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Appendix 2
dominant contribution of these sources to U.S. BC
emissions.
A2.2.1 Stationary Sources
Stationary sources in the NEI include both point
(fossil fuel combustion, industrial sources) and
nonpoint (biomass burning) source categories. The
basic method for estimating PM2.5 emissions for all of
these sources follows the simple conceptual formula
described in Equation 1:
E = AxEF(l-ER/100)
Where
(Equation 1)
• E = PM2.5 emissions (for example, in Tons);
• A = activity rate;
• EF = the emission factor, and
• ER = overall emissions reduction efficiency, %
Direct PM2.5 emissions are composed of both
filterable (solid) and condensable (gaseous) fractions.
The condensable fraction condenses rapidly in the
ambient air to form tiny liquid droplets. The sum
of the filterable and condensable fractions is what
is reported in the NEI for all source categories, and
these estimates are used in Chapter 4 of this report.
A2.2.2 Emission Factors
AP-42, Compilation of Air Pollutant Emission Factors
(EFs), has been published since 1972 as the primary
compilation of EPA's EF information. It contains
EFs and process information for more than 200 air
pollution source categories. More recently, AP-42 has
been transitioned into the FIRE 6.25 Data System,
which currently represents the most comprehensive
collection of emission factors (U.S. EPA, 2010d). It
currently contains thousands of records of (mostly)
filterable PM25 EFs updated through calendar year
2004.1
A source category is a specific industry sector or a
group of similar emitting sources. These EFs have
been developed and compiled from source test
data, material balance studies, and engineering
estimates. EFs can be as simple as an average rate
per unit process input. In most cases, EFs depend
on many variables such as process parameters,
effluent temperatures, ambient temperatures, wind
speed, and soil moisture. In these cases the formula
is applied to estimate emissions for a particular
set of conditions. Under some circumstances in
the inventorying of PM25, EFs and estimation
techniques are applied for analyses other than
those for which they were developed. The accuracy
and representativeness of the EFs are determined
by the reliability of the testing methodology,
how uniformly it is applied across sources, or the
engineering process information used to derive the
EFs.
EFs for some emissions categories are more
reliable than others. In some cases an EF may not
be available for a source category because of
insufficient or unacceptable data for generalization
across source type. Often it is difficult to determine
precisely what the certainty in the EF is. Thus,
the application of EFs requires subjectivity and
judgment from knowledgeable technical staff
for the application of concern. As discussed in a
previous chapter, users of EFs in national-, regional-,
and urban- scale studies should be cognizant of
their potential limitations, and other techniques
should be considered to improve the confidence in
PM emissions inventories. Several such approaches
have been developed and some continue to be
explored: continuous emission-monitoring sensors,
material balances, specialized source profiling for
composition and compositional material balances,
source sampling to obtain improved particle-size
distributions and location-specific emission rates,
near-source ambient characterization, and source
apportionment techniques. It is important to note
that the reliability of EF estimation decreases when
only a few source tests are used as the basis for the
factor, or when judgmental decisions are made from
analogy between technologies. Differences in EF
estimates also can develop if the current operations
or processes are significantly different from those
upon which the original EFs were derived.
When most people think of PM, they envision solid
particles that exist in the exhaust stream. However,
PM2.5 is composed of both a filterable fraction and
a condensable fraction (see earlier discussion). The
filterable fraction already exists in solid particle form
in the exhaust. The condensable fraction exists in
gaseous form in the exhaust stream but condenses
rapidly in the ambient air to form tiny liquid
droplets. Together, the filterable and condensable
fractions are referred to as direct emissions of PM2.5,
and the summed number is what is reported in
the NEI for all source categories. Most AP-42 EFs
do not quantify the condensable fraction of total
PM2.5 emissions. "Gap filling" techniques are used
to estimate condensable PM25 for many stationary
and area source categories. This introduces some
uncertainties in the emissions estimates.
1 http://cfpub.epa.gov/webfire.
272 Report to Congress on Black Carbon
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Black Carbon Emissions Inventory Methods and Comparisons
A2.2.3 Emissions Reduction Factors
The emissions-reduction factor (ER) in Equation 1
accounts for emissions controls employed on
a source. For example, these include various
effluent exit devices such as bag house filters and
electrostatic precipitators for removal of PM. Like
other process equipment, emissions controls have
variable operating performance depending on their
design, maintenance, and nature of the process
controlled. Thus, like EFs, values of ERs are overall
averages for specific processes and emissions-control
designs based on limited testing. Actual values of
ERs vary in time and by process in an undocumented
manner, adding significant uncertainty to emissions
estimates. Note that if no emissions controls are
applied, the abatement efficiency equals zero (ER=0)
and the emissions calculation is reduced to the
product of activity and the emission factor, EF.
A2.2.4 Activity Levels
The last piece of information needed in equation
1 to estimate PM2.5 emissions for sources is
activity patterns. Activity patterns (AP) describe
average temporal operating characteristics of a
process, including estimates of the down time for
maintenance or process failure. Values of AP for
point and non-point sources are each obtained in
different ways owing to the differing nature of the
sources.
Most point sources or industrial sources operate with
local permits, and these require information about
process emissions, including temporal characteristics.
For sources with CEMs for monitoring opacity
(roughly proportional to fine PM loading), such
as large utility boilers, real-time data are available
to derive activity patterns, and deduce emissions
variability over extended time periods. Further, point
sources keep and report records of output during
operating periods, and maintenance or other down
times.
There is a great deal of complexity in acquiring
activity data for nonpoint sources, which are
diverse in character, individually small, and often
intermittent, but collectively significant. Though
such sources are difficult to characterize, they are
generally important to PM emissions estimation
because their aggregated mass emissions can be
large and their chemical composition (e.g., BC) may
be important for estimating source attribution. One
good example of such a category is forest fires,
burning of land-clearing debris, and agricultural
burning.
Temporal resolution depends on the allocation
of emissions aggregated seasonally, weekly, daily
or by diurnal variation, depending on use and
industry activity patterns. The temporal allocations
allow for improved approximation of the actual
temporal patterns that can be important not only
for precise annual averaging using seasonal or daily
allocations, but also for short-term impairment
taken over periods of 24 hours or less. "Typical"
temporal variations for different sources have
been developed from surveys, activity analyses,
and expert consensus. These temporal models are
approximations that may deviate substantially from
actual emissions in a given location. Depending on
the requirements for precision in estimations, local
testing through observations and activity data may
be required, not only for large point or nonpoint
sources, but for smaller ones that may be of special
interest.
For nonpoint sources, emissions can be estimated
coarsely from "top-down" measures of activity at
the state or national level, including demographics,
land use, and economic activity. The construction
industry, for example, is based on the total annual
expenditures at the regional level. These estimates
are then allocated by county, using a procedure
linked with construction costs and estimated area
under construction. Because of their potential
importance as PM sources, considerable effort
has been devoted recently to the characterization
of emissions and activity patterns for non-point
sources. Another example is estimation of emissions
from fires, which depends upon knowledge of the
time, location, and areal extent of the burn, fuel
loading, types of combustible material and moisture
content. Recent efforts by EPA include the use
of process modeling and remote sensing data to
better estimate fire activity patterns and emissions
from fires (Blue Sky Framework, 2010). Finally,
residential wood burning is also an important local
source of PM and BC. Quantification of emissions
from this source category has been approached
through acquisition of data on how much fuel is
burned in fireplaces and woodstoves using national
consumption estimates. Where this source is a
large contributor to PM, local surveys of firewood
use are used to supplement and improve activity
level estimates. For all burning categories, the
PM emissions reported via AP-42 contain both
condensable and filterable emissions, so that the
uncertainties involved with arriving at total PM2.5 is
less compared to other point and non-point sources.
A2.2.5 Estimating BC and OC Emissions
Next, these PM2.5 emissions can be converted to
BC and OC by proper application of speciation
Report to Congress on Black Carbon 273
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Appendix 2
factors from the SPECIATE database. (See Appendix
1 for details on SPECIATE.) The equation used is
quite simple: PM2.5 Emissions (in tons) * fraction of
PM2.5 that is BC = BC emissions. This can be difficult
given that there are thousands of PM2.5 source
categories but only a limited number of speciation
profiles. Therefore, special attention must be given in
mapping specific profiles to source categories. These
details are explained in Reff et al. (2009). Application
of these methods to the inventories results in the 90
source categories for which BC and OC emissions are
reported in Chapter 4.
While the process for compiling BC emissions
inventories is reasonably straightforward, there are
important limitations in this process that introduce
uncertainties in final BC emissions estimates. These
include:
• The reliability of the PM2.5 emission factors used in
Equation 1. Some emission factors for point and
non-point sources are more reliable than others
(NARSTO, 2005).
• The reliability of condensable PM estimates
by source category. Some sources include PM
condensables as part of their testing protocol
(fires, residential wood combustion). Others do
not, and a generic emission factor (via AP-42) is
applied to estimate the amount of condensable
PM each source emits. This introduces a level of
uncertainty in determining final BC emissions
that is not currently accounted for. The source
measurements section of this report gives a
clearer indication of what the issues are and how
they can be improved.
• The reliability of activity levels used in Equation 1.
Some activity levels are generated using process
models (Blue Sky Framework, 2010), while some
are generated using surrogate information (U.S.
EPA, 2005a).
• Finally, many "augmentations" are done in the
emissions inventory processing steps. These
augmentations include scaling measured PM to
PM25 as well as assigning condensable emissions
estimates to point and nonpoint sources that
are not available via source testing. Some of the
impacts of the uncertainties in doing this have
been explored (NARSTO, 2005), but the issue has
not been dealt with holistically.
A2.2.6 Mobile Sources
In the U.S. inventory, mobile sources consist of the
following general categories of vehicles and engines:
• On-road gasoline, such as passenger cars and
light-duty trucks
• On-road diesel, including light-duty passenger
cars, light-duty trucks, and heavy-duty trucks.
Unlike in Europe, very few diesel passenger cars
are sold in the United States, making heavy-duty
diesel trucks the dominant vehicle type in this
category.
• Nonroad diesel, including construction,
agricultural, and other equipment
• Nonroad gasoline, including both 2-stroke and
4-stroke cycle engines such as those used in
lawn/garden equipment and recreational marine
• Commercial marine, classified by engine
displacement as categories Cl, C2, and C3
(ocean-going)
• Locomotives
• Aircraft, which are generally turbine aircraft rather
than the smaller piston gasoline-powered aircraft
BC emissions from on-road vehicles, both gasoline
and diesel, are now calculated directly using
EPA's new MOVES 2010 model. For other mobile
source categories, BC emissions are calculated
using methods similar to those described above
for stationary sources. EPA has released a number
of technical reports on MOVES. These give the
structure of the model including fleet and activity
data such as vehicle miles traveled, the default
national vehicle population, and vehicle activity
patterns.2 They also give information on exhaust
emission rates and deterioration by model year
for light-duty and heavy-duty vehicles.3 Similar
information is available on the EPA NONROAD
2 EPA (November 2010) Report EPA-420-R-10-026, MOVES 2010
Highway Vehicle Population and Activity Data, http://www.epa.gov/
otaq/modets/rnoves/420rlOQ26.pdf.
3 EPA (April 2001) Report 420-R-01-007 EPA's New Generation
Mobile Source Emissions Model: Initial Proposal and issues. MOVES,
http://www.epa.gov/otaq/rnodels/moves/rnovesback.htm.
EPA (February 2003) Report 420-R-03-005 Investigation for the
Physical Emission Rate Estimator to be Used in MOVES, http://www.
epa.gov/otaq/rnodels/moves/rnovesback.htm.
274
Report to Congress on Black Carbon
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Black Carbon Emissions Inventory Methods and Comparisons
model for things like equipment population,
emission factors, and engine turn over.4
The inventories given below for the United States
include all 50 states. They also include all marine
activity (both domestic and foreign) within 200
nautical miles of shore.
A2.2.6.1 On-road Gasoline and Diesel
For on-road gasoline and diesel vehicles, EPA's
emissions models directly calculate both total PM2.5
emissions and BC emissions. Recent improvements
in EPA's new MOVES 2010 model (U.S. EPA, 2010e)
as compared to the earlier MOBILE6.2 model (U.S.
EPA, 2003) include accounting for high emitters,
deterioration of PM emissions (i.e., increase in
PM mass) with higher mileage, and increased PM
emissions at lower temperatures.5 This model directly
calculates BC emissions (as well as other exhaust PM
components such as sulfates and OC), and accounts
for the significantly reduced BC fraction emitted
from on-road diesels due to application of diesel
particulate filters (DPFs) (required for heavy-duty
diesel trucks up to 80,000 pounds GVW beginning
with the 2007 model year). An important input for
the gasoline vehicle PM2.5 portion of the MOVES
model is a recent study examining PM emissions
from about 500 in-use vehicles (Coordinating
Research Council, 2008). MOVES accounts for the
lower EC/PM fraction (about 10%) for diesels with
DPFs. Several studies such as the Health Effects
Institute Coordinating Research Council study
(Khalek et al., 2009) have evaluated the EC/PM
fraction of heavy-duty diesel engines with DPFs
4 EPA (December 2005) Report EPA420-R-05-016 Exhaust Emission
Effects of Fuel Sulfur and Oxygen on Gasoline Nonroad Engines,
http://www.epa.gov/oms/nonrdmdl.htmfmodel.
EPA (July 2010) Report EPA-420R-10-017 Nonroad Engine Population
Estimates, http://www.epa.gov/oms/nonrdmdLhtmfmodel.
EPA (December 2005) EPA Report 420-R-05-017 Seasonal and
Monthly Activity Allocation Fractions for Nonroad Engine Emissions
Modeling, http://www.epa.gov/oms/nonrdmd[.htm#mode[.
EPA (July 2010) EP{A Report 420-R-10-015 Median Life, Annual
Activity, and Load Factor Values for Nonroad Engine Em issions
Modeling, http://www.epa.gov/oms/nonrdmd[.htm#mode[.
EPA (December 2005) EPA Report 420-R-05-018 Calculation of Age
Distributions in the Nonroad Model: Growth and Scrappage, http://
www.epa.gov/oms/nonrdmdLhtmfmodel.
EPA (July 2010) EPA Report 420-R-10-018 Exhaust and Crankcase
Emission Factors for Nonroad Engine Modeling - Compression
Ignition, http://www.epa.gov/oms/nonrdmdl.htmfmodel.
EPA (December 2005) EPA Report 420-R-05-021 Geographic
Allocation of Nonroad Engine Population Data to the State and
County Level, http://www.epa.gov/oms/nonrdmdl.htmfmodel.
5 MOVES also accounts for emissions changes with use of gasoline/
ethanol blends, although the effect on PM exhaust emissions from
use of gasoline/ethanol blends is extremely small, if not zero (U.S.
EPA, 2010e).
finding a fraction of about 10% versus the more
typical 70-80%. EPA plans to perform an uncertainty
analysis with MOVES, considering the various inputs
and what the range of uncertainty might be, and
what the overall uncertainty might be.
Gasoline OC and BC emissions increase dramatically
at lower ambient temperatures. To calculate this
increase for gasoline vehicles, we used calculations
done for EPA rulemaking packages for gasoline
PM, for which an hourly grid-cell temperature
adjustment was done as part of emissions
processing at the county level for each of the
over 3,200 counties. As a general rule, diesel PM
emissions are less sensitive to temperature for a
variety of reasons (lower importance of cold start
since many diesels trucks do not operate on short
trips; easier engine warm up since older diesels do
not have catalysts which take a finite time to warm
up during which emissions are higher). This means
that BC emissions from diesel vehicles are not
projected to increase as much at lower temperatures
as would be the case with gasoline vehicles.
MOVES can also be used to calculate tire and
brake wear PM2.5, with speciation factors applied
to calculate BC. Only a small fraction of the PM
from tire and brake wear is in the PM2.5 range, so
estimated BC emissions from these categories are
fairly small. However, a large fraction of tire wear PM
(about 22%) is BC. In the U.S. inventories reported
in Chapters 4 and 8, these detailed calculations at
the county level were done for 2005 and projection
years (2020, 2030) along with some less detailed
calculations (at the national level) for 1990. One
important thing to note is that the PM, BC, and OC
are relatively high from the on-road gasoline vehicle
fleet for 1990 due to the presence of a large number
of non-catalyst vehicles still remaining in the fleet.
A2.2.6.2 Nonroad Gasoline and Diesel
For nonroad engines (both gasoline and diesel
powered), EPA calculates BC emissions based
on PM emissions estimates from the NONROAD
model (U.S. EPA, 2008b). Also, the National Mobile
Inventory Model (NMIM) uses the current version of
the NONROAD model (NONROAD2008) to calculate
emissions inventories. The model incorporates
emission factors (in g/BHP-hr - that is, grams per
brake horsepower-hour), engine output (BHP-hr),
and usage data for a wide number of NONROAD
sources. For gasoline engines, 2-stroke cycle
engines are a separate category from 4-stroke
cycle engines. These engines have lubricating oil
mixed with the fuel so the exhaust VOC (and PM)
will be markedly different from that for the more
standard 4-stroke cycle engines. VOC from 2-stroke
Report to Congress on Black Carbon 275
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Appendix 2
cycle engines will contain higher molecular weight
compounds from the oil mixed with the gasoline;
oil compounds are typically in the C20-C37 range.
For these engines, the profile used to derive the
emissions estimates in Chapter 4 is that used for
non-catalyst equipped gasoline-powered motor
vehicles since these nonroad gasoline engines do
not have catalysts. This profile (92049) comes from
the EPA SPECIATE database and shows 10% of the
PM being BC. Admittedly, data specific for nonroad
gasoline engines, especially 2-stroke engines with
their oil combustion, are needed. EPA plans to add
NONROAD into MOVES at some point in the future.
Also, at some point possibly beyond that, EPA would
perform an uncertainty analysis on the nonroad
inventories.
BC emissions are then calculated by using speciation
factors denoting the percent of PM emissions
represented by BC. A speciation factor for nonroad
diesel engines not equipped with DPFs comes from
EPAs SPECIATE database (U.S. EPA, 2008a). The
profile used to derive the emissions estimates in
Chapter 4 (Profile 92035) is actually derived from
heavy-duty on-road diesels and has 77% of the PM
being BC. Beginning in calendar year 2012, many
if not most newly manufactured nonroad diesels
for that "model year" will be equipped with DPFs.
This technology reduces exhaust PM mass by over
90%, and the small amount of PM remaining has
relatively little BC. In effect, DPFs preferentially
reduce BC. Roughly 10% of the PM from a diesel
with a DPF consists of BC based on a large emissions
characterization program on four 2007 model
year on-road diesel truck engines equipped with
DPFs. The testing was done by Southwest Research
Institute for this program conducted by the
Coordinating Research Council (Khalek et al., 2009).
A critical factor in compiling BC inventories for
nonroad diesels is to correctly apportion the BC
emissions between pre-trap-equipped diesel
engines and trap-equipped diesels in any given
calendar year. The NONROAD model correctly
calculates the combined PM mass in a given calendar
year accounting for pre-trap and trap-equipped
diesels. Though it does not presently calculate
BC emissions separately, a later version of the
model under development will do so. Meanwhile,
when NONROAD is run, one can get a model year
emissions output for specific calendar years. One
can then probably manually take that model year
input and apply the higher BC speciation percent
(77%) to the pre-trap equipped engines and the
lower percentage (roughly 10%) to the new diesel
engines equipped with DPFs. The inventory numbers
presented account for this difference. For nonroad
gasoline, a speciation profile of 10% of the PM
being BC is used based on tests on older non-
catalyst light-duty vehicles. Most nonroad engines
do not have catalysts. Since almost no or limited
PM speciation has been obtained for the exhaust
of these engines, the most appropriate factor to
apply is based on older non-catalyst vehicles (of the
types produced before introduction of catalysts with
the 1975 model year). It is also important to note,
however, that 2-stroke cycle engine production will
be changed with the advent of new EPA emission
standards.
A2.2.6.3 Commercial Marine, Locomotive, and
Aircraft
Commercial marine, locomotive, and aircraft
emissions are calculated separately in spreadsheet
models, with separate BC speciation factors for Cl/
C2 commercial marine and C3 commercial marine
(the larger ocean-going vessels). For the smaller
vessels, the profile for nonroad diesel engines is
applied even though the higher sulfur content of
the fuel will lead to the PM containing higher sulfate
emissions than for nonroad diesels. DPFs will be
required for these vessels starting in 2014, reducing
the BC fraction to about 10% of the PM. However,
DPFs will only be used on some engine classes,
and implementation dates will vary (depending on
factors such an engine size). Thus, there is a need for
a model to correctly account for the implementation
of these standards. For now, a model year break-out
of PM emissions was done for both 2020 and 2030.
Separate BC/PM speciation factors were applied to
the PM emissions from the diesels with and without
DPFs. Currently, the diesel BC speciation factor of
77% BC/PM is used for C1/C2 commercial marine
for all years of analysis: 2005 as well as the non-DPF
equipped engines in 2020 and 2030. Evidence from
recent studies (Lack et al., 2009) suggests that a
lower BC speciation factor may be more appropriate
for C1/C2 marine.
PM emissions from C3 Marine have substantially
different PM speciation profiles than smaller diesel
engines used in C1/C2 Marine and on-road and
nonroad diesel. C3 marine diesels burn a high
molecular weight residual oil that contains very
high sulfur levels (up to 45,000 ppm versus the 15
ppm in on-road and nonroad diesel fuel). Past EPA
evaluations of C3 marine have used the EPA PM
SPECIATE profile of Residual Oil Combustion (U.S.
EPA, 2008c), which estimates a 1% BC speciation
factor.
For this report, an updated BC speciation profile was
estimated from studies available in the literature.
Results from relevant studies that measured BC
276 Report to Congress on Black Carbon
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Black Carbon Emissions Inventory Methods and Comparisons
Table A2-1. Summary of Recent Studies that Measured BCand PM Emission Rates from C3 Marine.
(Source: U.S. EPA)
Study Vessel Fuel Fuel Sulfur
y Content
Murphy etal. (2009)
Agrawaletal. (2008)
Petzold etal. (201 0)b
Lack etal. (2009)
Lack etal. (2009)
Post-Panamax Container
Suezmax Marine Tanker
Medium Speed Diesel Engine
Slow-speed diesel vessels'1
Medium-speed diesel vessels6
Heavy Fuel Oil
Heavy Fuel Oil
Heavy Fuel Oil
Variety
Variety
30,000 ppm
28,500 ppm
22,100 ppm
Variety"
Variety'
0.31%
0.50%
2.63%
7.33%
28.00%
a Lack et al. (2009) measured BC not EC with Light-Absorption Measurements (photoacoustic aerosol spectrometer
with adsorption at 532 nm), other studies measured EC using thermal-optical methods, but are referred to as BC for
comparison.
b Engine test, with engine load 85-110%.
c PM measurements come from 29 SSD ships; BC emissions come from 52 SSD ships.
ci Mostly high sulfur fuel (>5,000 ppm).
e PM measurements come from 12 vessels; BC emissions come from 51 vessels.
f Mostly low sulfur fuel (<5,000 ppm).
Table A2-2. Summary of Speciation Ratios of PM from Relevant Marine Studies. (Source: U.S. EPA)
Reference BC EC OC or OM BC Measurement EC/OC Measurement
Agrawal etal. (2008)
Murphy etal. (2008)
Petzhold etal. (2010)
(SSD) Lack etal. (2009)
(MSD) Lack etal. (2009)
(>5,000 ppm) Lack et al. (2009)
(<5,000 ppm) Lack et al. (2009)
0.85%
7.3%
28%
7.1%
51%
0.5%
0.3%
2.6%
11%
7.4%
21%
23%
16%
19%
35%
Multi-Angle Absorption
Photometer, MAAP
Photoacoustic Aerosol
Spectrometer with
adsorption at 532 nm
Thermal optical transmittance
(TOT) NIOSH 5040
Thermal-Optical Method.VDI 2465
Part 2
Organic Matter (OM) calculated by
subtracting S04,N03, and NH4from
aerosol mass spectrometer (AMS)
measurements
and PM emission rates from marine sources are
summarized in Table A2-1.
As noted in Table A2-1, there is substantial
variation in the reported BC/PM emission profiles
from these studies. Table A2-2 displays the BC,
EC, and OC speciation rates from the relevant
marine studies. Large variations between BC
(light-absorption measurements) and EC (thermal-
optical measurements) are observed for C3 Marine
emissions. Petzold et al. (2010) noted differences
in the EC and BC emissions up to a factor of 3 for
a medium speed diesel engine measured in the
laboratory. The discrepancies among different
BC emission factors for marine sources are
additionally noted in the literature (Petzold et al.,
2010). Considering the uncertainty of the values,
EPA selected a BC/PM speciation factor of 3%
which falls in the middle of the range of reported
values. Alternatively, one can also consider which
study has vessels that are most representative of
the C3 marine population or even a population-
weighted average. However, the differences in the
measurement techniques and definitions of BC
in the cited C3 Marine studies make it difficult to
combine the data across studies. EPA recognizes that
this is an area of active research and recommends
further work be conducted.
In Table A2-2, the results from the Lack et al. (2009)
study are subdivided according to vessel type: slow-
speed diesel (SSD) and medium speed diesel (MSD).
Lack et al. (2009) also grouped the BC emission
observations according to fuel sulfur content. Fifty-
one ship observations had low sulfur fuel content
(<5,000 ppm) and 42 ship observations of vessels
Report to Congress on Black Carbon
277
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Appendix 2
had fuel sulfur content greater than 5,000 ppm. From
the available data, BC and OC speciation factors were
calculated for each of the subcategories. The ships
in the Lack et al. (2009) study with low sulfur content
had much lower sulfate speciation factors and higher
BC speciation factors than the other studies.
To estimate the BC/PM factor for future years (2020
and 2030), the international fuel sulfur limits were
considered (Table A4-2) as well as the speciation
rates from the studies evaluated. Lowering the fuel
sulfur content is an effective method to reduce the
particulate sulfate, which comprises the majority
of the PM from marine vessels using heavy fuel oil.
Due to the substantial drop in fuel sulfur levels, the
BC speciation factor should rise in 2020 and 2030.
Due to limited data, the EPA chose a C3 marine BC
speciation factor of 6% for 2020 and 2030. For now,
EPA is choosing 11% as an OC/PM speciation fraction
for 1990 and 2005 with a higher fraction (58.6%) for
2020 and 2030 when fuel sulfur reductions occur,
especially in ECA areas.
Various inventories have also been prepared for
C3 marine using fuel consumption, emissions, and
vessel activity (Paxian et al., 2010).
For locomotives, as for C1/C2 marine, the HDDV
on-road profile (77% BC) is presently being used for
pre-2014 engines although available data suggest
this number might be too high. DPFs will be used
in subsequent years, reducing BC to about 10%. For
2020 and 2030, the PM model outputs are obtained
by calendar year and for the years when the
standards take effect, the 10% number is used.
For purposes of emissions inventory estimates,
aircraft operations are often broken into two basic
portions. The first portion, landing and take-off
(LTO) cycle, is normally defined to include aircraft
ground operations (taxi/idle) as well as aircraft
operations below 3,000 feet elevation in the local
airport terminal area. The second portion is referred
to as non-LTO that includes climb (above 3,000 feet)
to cruise altitude and descent from cruise to 3,000
feet. Together these portions comprise what is called
"full-flight" emissions.
Emissions for the LTO portion are fairly well
characterized. Engine emission rates are measured in
jet engine test cells during FAA certification testing;
it is believed that these measurements reasonably
predict engine emissions rates for aircraft in actual
LTO operations. Programs for evaluating and
controlling LTO emissions have been in place in the
United States for about thirty years. Today there are
LTO engine emissions standards for hydrocarbons,
carbon monoxide, oxides of nitrogen and smoke
number. While work is now underway to develop
a sampling and measurement procedure and
certification requirement for aircraft jet turbine
engine PM emissions, there are not yet specific
engine emission standards for PM. To address this
shortfall on at least an interim basis, FAA, working
with EPA, industry and academic experts developed
a methodology to estimate LTO PM emissions.
This methodology, known as the "First Order
Approximation" (FOA), uses information on smoke
number and other engine and fuel parameters to
estimate LTO PM emission rates for each engine
model (Kinsey and Wayson, 2007). This information
is then matched with airframe information on
number of engines to get a per LTO emission rate
for each aircraft type. Using the airport specific
information and the aircraft activity for each
airframe/ engine combination, the LTO PM Inventory
estimates are made. The total PM emission rate
includes all types of compounds contributing to
the PM mass. It is estimated that only about 13% of
the PM mass is BC; the remainder is composed of
sulfates and organics. The average BC PM emission
index (El) is in the range of 0.04-0.05 g/kg fuel
burned for the LTO portion.
The estimation of non-LTO BC emissions
depends on a very limited set of measurements.
Emissions testing in jet engine test cells does not
fully characterize PM emission rates at altitude
because they are conducted at sea level static
conditions and have to be carefully extrapolated
to altitude conditions, due to the differences in the
atmospheric environment and engine operating
conditions outside of the LTO—including cruise
altitudes. Although there are research models
available to estimate non-LTO BC, there is not yet
a consensus approach for estimating non-LTO PM
emissions as exists for LTO PM emissions. This is an
area of ongoing research within the scientific and
technical aviation communities.
However, two important points should be
recognized with regard to non-LTO PM BC
emissions. First, results from FAA's model entitled
"System for Assessing Aviation's Global Emissions"
(SAGE) indicate that total fuel burn during non-
LTO operations is about ten times that during the
LTO.6 Since the PM emission inventory is linked to
fuel burn, overall PM emissions during the non-LTO
portion of the "full flight" would be expected to be
larger than those during the LTO portion. Second,
this is an area of ongoing research and to-date
there are no less than six researchers who have
used various methods to estimate the El for PM BC
6 The FAA SAGE website is http://www.faa.gov/about/office_org/
headquarters_offices/apt/research/models/sage.
278 Report to Congress on Black Carbon
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Black Carbon Emissions Inventory Methods and Comparisons
Table A2-3. Estimates of Aircraft BC Emissions. (Source: U.S. EPA)
Reference Aircraft Engine(s) Measurement Condition BC El (g/kg fuel)
Kinseyetal. (2011)
Petzoldetal. (1999a)
Puescheletal. (1997)
Petzoldetal. (1999b)
Petzoldetal. (1999b)
Petzoldetal. (1999b)
Lilenfieldetal. (1995)
Anderson etal. (1998)
Dopelheuer(2001)
DCS
Various air
frames APEX 1
to 3
ATTAS
Concorde
B737-300
A31 0-300
VFW614
DCS
Multiple
n/a
CFM56-2C1
CFM56-7B24
CFM56-3B1/3B2
CFM56-3B1
RB211-535E4B
Rolls-Royce/SNECMA
M45HMk501
Olympus 593
CFM56-3B1
CF6-80C2A2
M45H
GE404
Multiple
CF6-50C2
Non-LTO thrust levels at Sea Level
Static (SLS)
Non-LTO thrust levels at SLS
In-flight
In-flight 16300m altitude
In-flight 7925 m altitude
In-flight 10670m altitude
In-flight 7925 m altitude
All thrust levels at SLS
In-flight; mass Els estimated from
number Els based on average
particle volume and mass
Modeled Cruise Simulation - DLR
Method using empirical calculation
0.021,0.026,0.032
0.028, 0.025
0.092
0.098
0.275
0.118-0.149
0.11-0.15
0.07-0.11
0.01
0.021
0.07-0.11
0.03-0.4
0.01-0.35
0.015
emissions during the non-LTO portion of "full flight"
(Lilenfield et al., 1995; Pueschel et al., 1997; Anderson
et al., 1998; Petzold et al., 1999a; 1999b; Dopelheuer,
2001; Kinsey et al., 2011). Some researchers have
used equivalent non-LTO thrust levels on the ground
while the estimates of others were based on in-
situ plume measurements of aircraft in flight. It is
difficult to make direct comparisons among these
values or to use this data to derive a point estimate
for the non-LTO BC El since they were developed
on different airframe/engine models of different
technology vintages using different measurement
approaches. While data from the published
researchers ranges from about 0.01-0.35 g/kg fuel
burned, the majority of the data lies in the range of
about 0.02-0.11 g/kg fuel burned. Each study has
its relative strengths and weaknesses and most of
the older engines with higher Els are no longer in
service. Table A2-3 summarizes the publicly available
literature on this issue.
A2.3 Development of International
Emissions Inventories for Black and
Organic Carbon
There are a number of methodological differences
between the approaches used to compile domestic
and international inventories. Specifically, in
contrast to EPA's method of using emission factors
paired with activity levels to estimate BC and OC
emissions, the most widely used global emissions
inventory described in Chapter 4 (Bond et al., 2004)
incorporates other factors to derive estimated BC
emissions, including fuel type, combustion source
technology type, and emissions controls. There is
extensive usage data on emissions from specific
vehicles and engines in the United States which is
used for input for EPA models. These data exists
to a lesser extent outside the United States. Fuel
consumption data, which are often used as a
substitute in global inventory calculations, are also
useful but do not have the detail that vehicle/engine
usage data have.
Like the U.S. inventory, global inventories typically
rely on Equation 1 (outlined above), for estimating
BC and OC emissions. That is, country-specific
emission factors are combined with appropriate
activity level information to yield an estimate
of emissions. For example, on-road cars, trucks,
buses, and all on-road mobile sources are generally
assessed through travel-based emission factors
and vehicle miles traveled (VMT). This approach
associates mobile source emissions with traffic
patterns, providing spatial and temporal information
about the distribution of emissions that can be
used in a variety of applications (for example, air
quality modeling). However, motor vehicle emission
factors are highly variable and uncertain because
of different vehicle types, ages, maintenance,
and operating conditions (Cadle et al., 2006). Fuel
composition data often can be obtained more
easily and accurately than activity measures such as
VMT. Fuel-based emission factors for fossil fuel and
bio-fuel combustion, for example, can be derived
easily from diluted in-plume measurements, using
Report to Congress on Black Carbon
279
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Appendix 2
simultaneous CO2 measurements to determine
dilution ratios and to relate other pollutants to the
weight of carbon in the consumed fuel (Chow et
al., 2010c). Fuel-based emission factors are very
common in global and large-scale inventories where
detailed information on source activity is very
limited.
Global BC and OC inventories are complicated by a
lack of specific, detailed information on source types,
emission factors, activities, and controls, especially
in the developing world. In such cases a simple
equation for calculating emissions based solely on
emission factors and activity levels cannot be applied
rigorously. Therefore, certain proxies have been used
to estimate BC and OC emissions.
A2.3.1 Specific Approaches Used in Global
Inventories
Most global inventories attempt to estimate BC
emissions, even though some of the emission
factors used seemingly represent testing on EC.
There are important differences in the way various
global estimates were generated, resulting in some
variation in the total emissions estimates generated
in different studies. It is useful to compare the
approaches in more detail.
The Bond inventories, which are the most extensively
used global inventories, were characterized in
Chapter 4. Bond etal. (2004) identified about 50
different combinations of fuel type and usage and
subdivided these into processes with different
emissions characteristics. This approach is based on
combining fuel composition data and assumptions
of combustion technologies and emissions controls,
and is very similar to earlier work done in the
literature (Klimont et al., 2002). Emissions for a fuel/
sector combination are calculated as an aggregate
of the contributions of all technologies within that
sector. The total emissions for each country, in turn,
are the sum across all fuel/sector combinations. The
reader is referred to published literature for more
details on the methods used and the uncertainties
inherent in their methodology (Bond et al., 2004;
2007). An overview of the Bond estimation procedure
is given in Figure A2-1. Using this method, global BC
emissions were estimated at about 8.9 million tons
per year, with an uncertainty range of 4.8 -24 million
tons/year. The United States accounted for about 6%
of global BC emissions in this inventory.
Other authors have compiled inventories based on
alternative methods. Penner et al. (1993) looked at
developing BC emission in two ways: first, based on
fuel consumption estimates, and second based on
BC/SO2 ratios. In examining the relationship between
ambient BC and SO2 concentrations in urban
areas around the world, the authors found strong
correlations in source areas and also that various
sources had characteristic BC/SO2 ratios. Site-specific
BC/SO2 ratios were transformed to BC emissions
using available SO2 emissions estimates for each
country/world region. The result was a global BC
emissions estimate of about 26 million tons per
year from urban fuel use. Penner et al. (1993) also
calculated global BC emissions on the basis of fuel
consumption, assuming constant emission factors
for commercial and domestic coal, diesel fuel, wood,
and bagasse combustion, yielding a total of 14
million tons BC/yr, with 7.3 million tons/year from
fossil fuel combustion and 6.7 million tons/year
from wood and bagasse fuel burning. Even though
global estimates from the two methods differed by
a factor of about 2, larger differences were found for
individual countries and regions.
Cooke and Wilson (1996) compiled published
estimates for biomass areal density, amounts above
ground, the fractions burned, and emission factors
for different fuel types (e.g., forests, savanna).
Agricultural burning and biomass combustion for
heating and cooking were not included. Country-
specific fuel consumption rates were compiled
for industrial, domestic, and combined sectors
for solid, liquid, and gaseous fuels. Country-wide
IEA fuel
consumption data
Country/fuel/sector
divisions from IEA
Regional fuel use by
combustion type
Emission factors,
size information,
speciation
Country estimates
of BC/OC emissions
Urban/rural population
Land cover
Gridded total
emissions
Figure A2-1. Bond et al. Methodology for Developing
Emissions Estimates. (Source: Bond et al., 2004)
280
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Black Carbon Emissions Inventory Methods and Comparisons
emissions were distributed to grids according to
population density. The Cooke and Wilson global BC
inventory of about 15 million tons was comprised
of 9 million tons and 6 million tons from fossil fuel
and biomass combustion, respectively. The fossil
fuel component of 9 million tons was approximately
one-third that estimated by Penner using the BC/
SO2 ratio approach (i.e. 26 million tons), but similar
to emissions based on Penner's fuel consumption
approach (i.e., 7.3 million tons from fossil fuel
combustion).
Liousse et al. (1996) reported global BC and OM
(organic mass) emissions. OM is divided by 1.3
and converted to measured elements in organic
molecules. Their inventory includes categories for
biomass (i.e., savanna and forest fires), agricultural
waste, wood fuel, and dung combustion as well as
domestic coal and diesel fuel combustion. Global
fossil fuel combustion (7.3 million tons BC /year)
and biomass burning (6.2 million tons EC /year) total
13.5 million tons EC/year, lower than the estimate
of 15 million tons BC/year from Cooke and Wilson
that excluded agricultural burning and biomass
combustion for fuel and energy production. Louisse
et al. (1996) also estimated global OC emissions of 69
million tons/year, with 24 million tons OC/year from
fossil fuel and 45 million tons/year from biomass
burning.
Then, Cooke et al. (1999) refined Cooke and Wilson
(1996) by considering the relative ages of vehicles
in developed and developing countries and particle
size differences for controlled and uncontrolled
combustion processes. The estimated global BC
inventory from fossil fuel combustion of 7 million
tons in Cooke et al. (1999) was consistent with
estimates of 9 million tons BC/year by Cooke and
Wilson (1996) and 7.3 million tons BC/year by Louisse
et al. (1996). The Cooke et al. (1999) estimate of
global OC emissions was about 11 million tons,
about half that of Louisse et al. (1996) (24 million
tons OC/year).
Cofala et al. (2007) have used a global version
of the Regional Air Pollution Information and
Simulation (RAINS) model to estimate anthropogenic
emissions of BC and OC (along with numerous
other pollutants). The authors rely on the RAINS
methodology (Klimont et al., 2002) for particle
emissions, which they modify to capture regional
and country-specific characteristics of BC and OC
emissions as laid out in other references (Kupiainen
and Klimont, 2007) and extend it to developing
regions with data from Bond et al. (2004). Their
methods result in an estimate of 6 million tons of BC
emitted globally in 1995 and about 5.9 million tons
of BC emitted in 2000.
A2.3.2 Specific Approaches Used in
Regional Inventories
Chapter 4 also discusses the information available
from alternative inventories available for specific
countries or world regions. Table A2-4 provides a
comparison of key differences in data and methods
between some of these inventories.
A comparison among the regional inventories listed
in Table A2-4 yields some interesting information. It
is also helpful to compare the emissions estimates
from these inventories with estimates in the
appropriate portion(s) of the Bond/Streets global
inventory. Specifically:
• Cao et al. (2006) used emission factors from
Cooke's 1999 global inventory and Streets'
inventory of China BC emissions for the year
2000, along with Andreae and Merlet (2001)
emission factors for biofuels. They developed a
local inventory based on specific emission factors
for crop straw used in cooking stoves by testing
five different types of straw in a combustion
tower designed to simulate Chinese cooking
stoves. Their national BC emissions estimate
for 2000 in China is about 1.7 tons, which is
somewhat higher than the Bond estimate of
1.6 million tons and the Streets estimate of
1.2 million tons. The authors attributed this
to their inclusion of coal combustion in rural
industry and rural residential sources, which they
noted are often underestimated in more global
estimates. They outlined the residential and
industrial sectors as being the most important in
contributing to Chinese BC emissions.
• Streets et al. (2001) measured Chinese BC
emissions for 1995, using mostly emission
factors from other literature sources. Their
inventory focuses on submicron BC emissions
rather than bulk emissions, because submicron
emissions are more relevant to radiative transfer
calculations. Their study noted also that most
inventories assume that the fraction of BC that
makes up PM2.5 remains constant throughout the
combustion process. Streets et al. (2001) states
that smoldering combustion, while releasing a
higher amount of particulate matter, does not
exhibit temperatures high enough to produce the
same proportion of BC. They propose a negative
correlation between particulate emissions and
the fraction of BC emitted. Other observations
by this study were that removal efficiency of
particles for the industrial sector is lower and
less documented than that of the power sector,
making emissions from the Chinese industrial
sector more uncertain and variable, and that
Report to Congress on Black Carbon 281
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Appendix 2
Table A2-4. Regional Inventories of BC and OC Emissions. (Source: U.S. EPA)
Carbon Emission
Inventory
Reference
Region,
Base Year
(Resolution, if
available)
Emission Source
Categories
Source of
Emission Factors
Source(s) of Activity
Information"
Reddy and
Venkataraman
(2002b, Fossil
Fuel) Reddy and
Venkataraman
(2002a, Biomass
Combustion) Black
Carbon and Organic
Carbon
Streets etal. (2003)
Black Carbon Only
Cao etal. (2006) Black
Carbon and Organic
Carbon
Streets etal. (2001)
Black Carbon
Pa rashar etal. (2005)
Black Carbon and
Organic Carbon
Sahu etal. (2008)
Black Carbon
Dickerson etal.
(2002) Black Carbon
India, 1996-1997
(.25 x .25)
Asia, 2000
(1x1 to . 08 x. 08)
China, 2000
(.2x.2)
China, 1995
India, 1995
India, 2001
(1x1)
Asia, 1991 and
2001
4 utilities
5 coal combustion
8 industrial
2 residential/ commercial
8 transportation
4 biomass/biofuels burning
Each of the 22 Asia countries
(plus international shipping)
has power generation,
industry, and domestic
sectors divided into 3
categories (i.e. coal, oil,
or biofuel.and other), 10
transportation categories,
and 3 biomass burning
categories.
Includes 5 sectors (i.e.
power generation, industry,
residential, transportation,
and biomass burning)
separated by 363 large
point (including 285 power
plants) and area sources (e.g.
population, gross domestic
product) with 18 different
sector-fuel type combination.
Covers 37 different source
types over five sectors
(power, industry, residential,
transport, field combustion)
and 13 different fuel types,
including biofuels
Fossil fuel, biofuel.and
biomass burning
Categorized into area sources
and IPS and then by fuel
type
Same as 37 different source
types identified in Streets et
al. (2001)
Literature review
U.S. EPAAP-42
Customized
emission factors
to fit Indian
technology
Literature review
U.S. EPAAP-42
MOBILES model
Literature review
Laboratory tests of
biofuel emissions
from cooking
stoves
Literature Review
Literature Review
Laboratory tests of
biofuels and soft
coke
Literature review
Literature Review
MOBILE 5, also
used CO/BC
ratio to estimate
emissions
Fossil fuel consumption from Central
Board of Irrigation and Power,
Cement Manufacturers' Association,
Centre for Monitoring of India
Economy, The Fertilizer Association
of India, Ministry of Coal, Ministry of
Industry, and Ministry of Petroleum
and Natural Gas, Statistics for Iron
and Steel Industry in India. Biofuel
consumption in rural and urban from
Tata Research Institute and National
Sample Survey and Forest coverage
from the Forest Survey of India.
RAINS-Asia simulation (2000 forecast
from the 1995 base year), except
for China which was compiled
on a provincial basis. For the
transportation sector, used World
Road Statistics and World Motor
Vehicle Data.
Point source activity from State
Power Corporation of China and
Editorial Board of China Rural Energy
Yearbook. Area sources activity
from National Bureau of Statistics
and various government agencies,
mainly at the county level.
Fuel consumption data by sector and
fuel type were developed within the
framework of the RAINS-Asia model
(Hordijketal., 1995). Generated by
China's Energy Research Institute.
Fossil fuel consumption from TEDDY
2001/2002, biofuel use and biomass
burning taken from Reddy and
Venkataraman (2002b).
Activity data collected from Central
Electricity Authority (CEA), Census
of India, Ministry of Coal, Ministry
of Road Transport and Highways,
Ministry of Agriculture.
RAINS-Asia model, Tata Energy
Research Institute.
282
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Black Carbon Emissions Inventory Methods and Comparisons
Carbon Emission
Inventory
Reference
Region,
Base Year
(Resolution, if
available)
Emission Source
Categories
Source of
Emission Factors
Source(s) of Activity
Information"
Lamarque etal.
(2010) Black Carbon
and Organic Carbon
Mitra etal. (2002)
Black Carbon and
Organic Carbon
Mayol-Bracero etal.
(2002) Black Carbon
Ohara etal. (2007)
Black Carbon and
Organic Carbon
Derwent etal. (2001)
Global, 2000
(.Sx.5)
India, 1996
India, 2000
Asia, 2000 (.5 x. 5)
Western
Europe,1995-1998
12 different sectors over 40
different regions
7 different fuel types
8 different fuel types across
four (five?) sectors
Four sectors broken down
into fuel type (coal, oil,
biofuel, others)
All sectors
Literature Review
(Junker and
Liousse, 2008
combination
emission factors)
Literature Review
(from Cookeand
Wilson, 1996)
Literature Review
(Streets etal.,
2001))
Literature Review
U.S. EPAAP-42
Back calculated
using dispersion
modeling and
ambient data
Biomassfrom RETRO, GICC, and
GFEDv2 inventories; ship data from
International Maritime Organization
(IMO); aircraft data from AER02K
database, EUROCONTROL, Bond et
al. (2007), and Junker and Liousse
(2008).
RAINS-Asia model.
LPS activity data from China State
Grid Company, RAINS-Asia, Fuel
consumption from International
Energy Agency (IEA) or UN Energy
Statistics Yearbook. Biofuel
consumption from RAINS-Asia.
NA
i Unless otherwise noted, refer to main Emission Inventory Reference for a list of specific publications and databases from which underlying
data were drawn.
domestic emissions in China are responsible
for over 80% of Chinese BC emissions. The final
estimate of BC emissions was about 1.5 million
tons, which is lower than the Bond estimate of
1.6 tons but higher than emissions estimated via
other regional Asian studies REAS and TRACE-P
(1.3 million tons and 1.0 million tons, respectively).
Reddy and Venkataraman (2002a; 2002b)
estimated BC emissions in India for the year 1996
by developing emission factors using Indian fuel
composition and indigenous pollution control
technology. Emission factors of coal for the
power and industrial sector were derived from
those of the EPA. Domestic emission factors
were taken from Indian literature sources (Gray,
1986). Transportation factors were taken from an
average of PM emission factors from countries
with similar transportation statistics. The study
noted that most of the BC emissions from India
were from the transportation sector (almost
60%). Overall they predicted BC emissions from
1996-1997 to be 0.36 million tons annually. This
estimate is lower than almost all other estimates
of Indian BC emissions. The authors claimed their
lower estimate was due to different emission
factors - other emission factors used in more
global inventories were too high due to improper
differentiation of fuel composition, combustion
type, and PM2.5 composition differences.
Parashar et al. (2005) estimated Indian BC
emissions for 1995 and used Bond emission
factors for fossil fuel combustion. Biomass
combustion emission factors were determined by
actually combusting different types of fuels in a
U shaped chimney. They found that dung cakes
released a particularly high amount of particles
due to a smoldering combustion, which releases
more particles than other types of burning. Their
final emissions estimate for India was 0.92 million
tons BC, which is higher than Bond's estimate
of 0.64 million tons of BC. The higher emissions
could be due to the higher emission factor they
found for dung cakes, which accounted for a
higher than proportional amount of domestic
emissions.
Mitra et al. (2002) estimated Indian BC emissions
for the year 1996 using Cooke's 1999 emission
factors for "under developed" countries. They
only calculated emissions for four fossil fuel
categories: coal, diesel, gasoline, kerosene. Their
annual emission estimate was approximately
Report to Congress on Black Carbon
283
-------
Appendix 2
1.1 million tons of BC. This is much higher than
Bond's estimate of 0.63 million tons BC in India.
The higher emissions could be due to the use of
emission factors for under developed countries, in
that Bond et al. in their estimates may have used
emission factors from more developed countries
to represent fossil fuel combustion characteristics
in India.
Sahu et al. (2008) estimated BC emissions
from India for the years 1991 and 2001. They
used Cooke's 1999 emission factors for "under
developed countries" for fossil fuel combustion
and Reddy's regionalized 2002 emission factors
for biofuel combustion. Their final estimate of BC
emission in India was about 1.5 million tons per
year, higher than any other inventory despite the
fact that they did not inventory small industry.
Their high estimates could be due to using under-
developed country emission factors for all fossil
fuel combustion sectors and also from use of
diesel activity information which did not represent
current conditions.
Mayol-Bracero et al. (2002) calculated BC
emissions in India using the Chinese emission
factors developed by Streets et al. (2001) and
national activity data from the GAINS model.
Their final estimate of BC emissions in India (2000)
was 0.5 million tons, slightly lower than Bond's
estimate of about 0.6 million tons.
Ohara et al. (2007) developed REAS (Regional
Emission Inventory for many parts of Asia) for
several pollutants including black carbon for the
period 1980 - 2003. Emissions were calculated
as a product of activity data, emission factors,
and removal efficiency of controls. BC emission
factors were taken from Streets et al. (2003a),
and characterized into developed countries and
countries with no known emission controls (this
category included India and China). The emission
factors for developed countries changed several
times over the time period of the inventory.
Chinese BC emissions in 2000 were 1.2 million
tons and Indian emissions were estimated to be
about 0.9 million tons. These numbers compare
fairly well to other estimates for those countries.
The inventory noted the domestic sector as the
main contributor to BC emissions.
Dickerson et al. (2002) used two different
approaches to measuring BC emissions in India
and other South Asian countries. They first did a
bottom up inventory using emission factors and
activity level data. They assumed that South Asia
source types of BC were similar to those of China,
and they obtained energy use information from
the RAINS-Asia model. For residential biofuel
combustion, they used the emission factor 1 g/
kg, which was taken from measurements in the
literature (Muhlbaier-Dasch, 1982) and was similar
to that used in the Reddy and Venkataraman
study outlined earlier. Their estimates of BC
emissions from power plants were lower than
Reddy and Venkataraman because of smaller
emission factors due to a high level of ash in
the particulate emissions. Indian vehicles were
assumed to be similar to Chinese vehicles; the
authors used emission factors from Streets'
2001 work. The final BC estimate for India was
0.56 million tons. Their estimates differ from
Penner et al. (1993), Cooke and Wilson (1996),
and Cooke and Wilson (1996) as well as Bond et
al. (2004), because of possible inclusion of ash
in the emission factors, omission of biofuels,
and difference in time periods. For the second
method, CO emissions were used as surrogates to
estimate BC emissions. They found that that total
BC emissions for India using this method were
more on the order of 2 million tons. The team
concluded that bottom up inventory estimates
produce much smaller values of BC emissions
than do actual in field observations (or "top
down" estimates), which could imply errors in
calculating these inventories "bottom up."
Zhang et al. (2009) focus on the INTEX-B mission,
the goals of which were to quantify transport and
evolution of Asian pollution to North America
and assess its implications for regional air quality
and climate. The inventory improved China's
emissions estimates by balancing the spread
of new and old technology in China's industrial
sector and improving energy statistics. For other
Asian countries, the mission used IEA energy
statistics and emission factors documented in
Klimont et al. (2002). The INTEX-B mission also
incorporated inventories that were thought to
be more accurate representations of individual
countries, such as the Indian inventory from
Reddy et al. (2002a; 2002b), the Japan inventory
from Kannari et al. (2008), and the South Korean
inventory from Park et al. (2001). China emissions
were 1811 Gg and India emissions were 344 Gg
for 2006. INTEX-B also included small industry
emissions, but noted that they were uncertain
of the numbers. The authors also noted that
for Southeast Asia, the activity level data was
extrapolated and there were few local emission
factors, so the data may not be very accurate.
This mission was seen as an improvement on
the previous TRACE-P due to the updated
methodology and collaboration with local
inventory efforts.
284 Report to Congress on Black Carbon
-------
Appendix 3
Studies Estimating Global and
Regional Health Benefits of
Reductions in Black Carbon
Geographic Results of Study Mitigation Measures Pollutants Reference
Studies of Mitigation Strategies for Ambient Reductions in BC
Global, Arctic
Global
Global
Global
Global
Fossil fuel soot (FS) and biofuel soot and
gases (BSG) contribute to global warming,
with FS being the greater contributor per
unit mass. However, BSG may contribute 8
times more in premature mortalities than
FS due to greater population exposures
to BSG.
Avoid 240,000 annual premature
mortalities in China, 30,000 elsewhere
globally. Find reductions in sulfates, OC,
and BC collectively lead to loss in net
negative radiative forcing.
Relative to no extra controls, imposing
tighter vehicle emission standards in
developing countries avoids 120,000-
280,000 premature air-pollution related
deaths in 2030.
Halving global anthropogenic BC
emissions avoids 157,000 premature
deaths annually worldwide, the vast
majority of which occur within the source
region. Most of the avoided deaths are
achieved by halving East Asian emissions,
but South Asian emissions have 50%
greater mortality impacts per unit BC
emitted than East Asian emissions.
Residential and industrial emissions
contribute disproportionately to
mortality due to co-location with global
population. About 8 times more avoided
deaths estimated when anthropogenic
BC+OC emissions halved compared with
halving BC alone.
Implementing all measures would
avoid 1-5 million PM2.5and 03-related
premature deaths annually based on
2030 population, with the vast majority
achieved by BC emissions controls. About
80% of the avoided deaths occur in
Asia. Avoided deaths occur regardless
of simultaneous implementation of low-
carbon C02 measures.
Elimination of global
anthropogenic FS and BSG.
50% reduction in China's 2030
S02, OC, BC emissions from 2000
levels.
Imposition of tighter vehicle
emission standards (e.g. Euro 6
standards for light-duty vehicles)
in China, India, Africa, Middle
East, Brazil, and the rest of Latin
America by 2030.
50% reductions in anthropogenic
BC emissions globally, from
8 major world regions, and
from 3 major economic
sectors (residential, industrial,
transportation).
Suite of methane mitigation
measures, "technical" BC
mitigation measures (ex.
improving coke ovens and
brick kilns and increasing use
of diesel particulate filters), and
"non-technical" BC mitigation
measures (ex. eliminating high-
emitting vehicles, banning open
burning of agricultural waste, and
eliminating biomass cook stoves
in developing countries).
PM2.5from fossil
fuel soot, biofuel
soot, & methane
S02, OC, BC
S02, OC, BC, N03,
03
BC.OC
S02, OC, BC, N03,
03
Jacobson
(2010)
Saikawa etal.
(2009)
Shindelletal.
(2011)
Anenberg etal.
(2011)
UNEPand
WMO(2011a)
Report to Congress on Black Carbon
285
-------
Appendix 3
Geographic Results of Study Mitigation Measures Pollutants Reference
Studies of Indoor and Ambient Mitigation Strategies for BC
China, India,
Africa
China, India
Benefits of mitigation exceeded costs by
factors of 3.6 to 13.6 to one.
Find BC mitigation strategies involving
indoor BC stove emissions and diesel
BC emissions reductions in urban cities
are win-win opportunity for climate and
public health.
Improved stoves in China and
India for domestic heating and
cooking, coal to briquette use for
domestic cooking and heating,
and community forestry programs
to control savannah and open
burning in Africa.
Indoor reduction of BC from
replacement of stoves used for
cooking and home heating, and
strategies to reduce BC emissions
from diesel vehicles used in urban
cities.
BC, OC, S02
BC, OC, S02
Baron etal.
(2009)
Kandlikar etal.
(2009)
Studies of Indoor Mitigation Strategies to Reduce BC
China
India and UK
Climate and human health benefits to
cost ratio of 6 with about 69% of these
benefits associated with human health.
Low emission stoves in India result in
12,500 fewer DALYs annually and energy
efficiency in the UK households results in
850 fewer DALYs per year.
Household fuel intervention.
Energy efficiency in UK household
heating and 150 million low-
emission cookstoves in India.
BC, OC, S02
BC, OC.sulfates
Smith etal.
(2008)
Wilkinson etal.
(2009)
286
Report to Congress on Black Carbon
-------
Appendix 4
Efforts to Limit Diesel Fuel Sulfur
Levels
As discussed in Chapter 8, the availability of low-
sulfur diesel fuel is imperative for many emissions
control strategies. Sulfur in fuel will poison the
catalysts that are built into passive DPFs, thus
rendering them ineffective. DPFs work ideally
with 50 ppm or less sulfur diesel fuel ("low-sulfur
diesel"). Thus, nations that have adopted low
sulfur requirements for diesel fuel of 50 ppm or
less are best positioned to adopt more stringent
emission standards for new motor vehicles, and
have more flexibility to target emissions from in-
use vehicles. Nations with established standards
of 500 ppm or less have more limited institutional
and technological potential for further reductions.
Nations with nominal or no limits on sulfur in diesel
fuel are unable to adopt technology-based standards
or controls on in-use engines that would offer
significant reductions in BC.
Aside from the United States, Canada, Japan, and
the European Union, 50 ppm or less sulfur diesel fuel
is not common. Only a few metropolitan areas in
developing Asia have 50 ppm sulfur diesel available
(USAID, 2010a). However, several countries around
the world have adopted schedules that require the
use of lower sulfur diesel fuel between 2010 and
2015:
• Africa: Morocco established limits of 50 ppm in
2009, and Tunisia will require 50 ppm fuel in 2014-
2015.
• Americas and Caribbean: Mexico adopted ULSD
(< 15 ppm) in 2009 nationwide, while Chile
and Brazil have mandated ULSD in urban areas
between 2009 and 2013. Several other nations
have established requirements for diesel fuel with
50 ppm sulfur, either nationwide (Columbia 2013,
Chile 2010, Uruguay 2010) or in large urban areas
(Argentina 2012, Colombia 2010).
• Caucasus and Central Asia: Armenia and
Kazakhstan both introduced requirements for 10
ppm diesel fuel in 2010. Georgia adopted national
standards for 50 ppm diesel fuel in 2010.
• East Asia and Pacific Islands: Malaysia required
50 ppm diesel fuel in 2010 and is requiring
10 ppm diesel fuel in 2015. Singapore, Malaysia,
and the Republic of Korea have established
national sulfur standards of 50 ppm in diesel
fuel between 2007 and 2010. Thailand is limiting
diesel fuel to 50 ppm sulfur in 2012. Malaysia and
the Republic of Korea plan to adopt 10-15 ppm
sulfur limits between 2010 and 2015.
• Eastern Europe: Ibania and Belarus plan to
require 10 ppm sulfur in diesel fuel in 2011-
2012. Croatia, Russia, and Turkey have adopted
standards of 50 ppm between 2008 and 2010,
though numerous fuel grades continue to be
sold.
• South Asia: China limits diesel sulfur to 50 ppm in
Beijing (2008), Hong Kong, and Macao; diesel fuel
in Taiwan is limited to 50 ppm sulfur after 2005
and 10 ppm starting in 2011. For selected urban
areas, India is requiring the use of 50 ppm sulfur
diesel fuel in 2010.
• Southwest Asia/Middle East: Israel required
10 ppm sulfur in diesel fuel in 2009, while Qatar
is requiring it in 2012. Saudi Arabia and Syria will
require 50 ppm fuel in 2014-2015.
Numerous other countries have established diesel
sulfur limits of 500 ppm prior to 2015, including
Azerbaijan, Brazil (outside urban areas), Ecuador,
Fiji, India, Malawi, Mozambique, Oman, Pakistan,
the Philippines, South Africa, Sri Lanka, Thailand,
Vietnam, and Zimbabwe.
Among nations with less stringent standards on
fuel sulfur (e.g., 2,000-10,000 ppm) in either all or
part of their territory, some have lowered the limits
in recent years. For example, outside urban areas,
Argentina and Peru are reducing allowable sulfur
to 1500 ppm between 2010 and 2012, from levels
of 2500-3000 ppm introduced in 2006. Venezuela
reduced allowable sulfur from a standard of 5,000
ppm established to a new standard of 2,000 ppm in
2010. Notable among nations of sub-Saharan Africa,
Mauritius established a diesel fuel sulfur standard of
2500 in 2001. Moving to lower sulfur levels in these
regions is hampered by economic and technical
barriers.
Report to Congress on Black Carbon
287
-------
Appendix 4
Among nations without sulfur standards, some
include oil producing nations, such as Egypt, Iran,
and Kuwait. Many sub-Saharan African nations lack
national sulfur standards. In the former Soviet Union,
many central Asian countries base their national
standards on Russia's GOST 305/82 standard for
diesel fuel (2,000 ppm). Nevertheless, some nations
have diesel fuel with sulfur levels that meet the
national standards of countries from which they
export. For example, diesel fuel in Lesotho, Namibia,
Swaziland, and Botswana meets the 500 ppm
national standard established in South Africa, from
which they import their fuel.
Through the Partnership for Clean Fuels and
Vehicles (PCFV) (http://www.unep.org/transport/
pcfv/), UNEP continues to work with developing
nations to identify opportunities and build capacity
to establish lower sulfur levels. For example, the
PCFV holds workshops in Africa, Asia, and the
Americas, gathering local scientists, engineers, and
officials to discuss scientific evidence and economic
impacts of how diesel fuel sulfur levels affect cities
in developing countries. These meetings follow on
PCFV's successful campaign to eliminate lead in
gasoline, which recently celebrated the complete
phase-out of lead in African gasoline.
Several regional intergovernmental agreements have
also been signed by representatives at the ministerial
level. In February 2008, environmental ministerial
officials from Latin America and the Caribbean in
Santo Domingo, Dominican Republic agreed to
promote sulfur reduction in fuel throughout the
region, with a target goal of 50 ppm. In July 2009,
several west and central African environmental
ministers signed a regional framework agreement
on air pollution, including goals to adopt 3500 ppm
fuel sulfur limits by the end of 2011, with a goal
of 50 ppm fuel by 2020. Though non-binding on
governments, these agreements suggest that there
is significant impetus to reduce sulfur levels in fuels
used in the developing world.
In addition to governmental and intergovernmental
efforts to reduce diesel fuel sulfur levels, several
private sector initiatives also exist. Vehicle industries
around the world have recognized the value of
reduced sulfur for enabling lower-emissions vehicles
and high-efficiency combustion technologies. In
2002, vehicle and engine manufacturers from the
United States, Europe, and Japan published a report
on worldwide fuels harmonization, which promoted
lower sulfur levels in gasoline and diesel fuel.
More recently, the African Refiners Association has
developed a set of "AFRI" fuel specifications (AFRI-1
through AFRI-4) as a developmental pathway for
African development of <50 ppm sulfur.
Table A4-1 gives recent information on national
standards for on-road diesel sulfur limits, and
estimates of current sulfur levels. In addition to the
efforts described above, Chapter 8 also mentions
the limits on sulfur content of marine fuel being
phased in under requirements from the IMO.
Table A4-2 provides details regarding the fuel
sulfur levels allowed for C3 marine fuel within ECAs
and globally outside of ECAs, and the schedule
for phase-in of tighter limits on sulfur content of
this fuel. For this table, the Global and ECA fuel
standards shown are the maximum fuel sulfur levels
allowed under MARPOL Annex VI for ships with
engines over 130 kW.1 The date on which the ECA
requirements become enforceable for a specific
geographic area depends on the date the treaty
amendment incorporating the ECA enters into force.
1 MARPOL is an abbreviation of "marine pollution," and is the
acronym used to refer to the International Convention on the
Prevention of Pollution from Ships.
288
Report to Congress on Black Carbon
-------
Table A4-1. International Regulations and International Agreements on Diesel Fuel Sulfur Levels (in ppm). (Source: U.S. EPA)
Country
2011 2012 2013 2014 2015
2016 2017 2018 2019
Americas and Caribbean
Mexico
(National)
Mexico
(Northern)
Argentina
(Urban)
Argentina (Non-
Urban)
Barbados
Bolivia
Brazil (Non-
Urban)
Brazil
(Metropolitan)
Chile (National)
Chile (Santiago)
Colombia
(National)
Colombia
(Bogota)
Costa Rica
Cuba
Dominican
Republic
Ecuador
(National)
Ecuador (Urban)
Ecuador (Non-
Urban)
El Salvador
Guatemala
Honduras
Panama
Peru (Urban)
?
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
500
300
1500
2500
500
15
1500
2500
500
15
1500
2500
15
15
1500
2500
15
15
(500)
2500
(500)
2500
50
1500
No existing or planned standards.
7
7
7
7
7
7
7
7
3500
2000
7
7
7
7
7
3500
2000
7
7
7
7
7
3500
2000
7
7
7
7
7
3500
2000
7
7
7
7
7
3500
500
7
7
7
7
7
3500
500
350
50
2500
7
7
3500
500
350
50
2500
7
?
3500
500
350
50
2500
?
?
3500
50
350
50
2500
?
(2000)
1800
50
50
10
500
50
1800
50
1800
50
1800
10
50
500
No existing or planned standards.
No existing or planned standards.
No existing or planned standards.
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
5000
7000
7
7
7
7
1500
(500)
7
7
7
7
1500
500
7
?
7
7
1500
500
7
?
7
7
1500
50
500
(500)
(500)
(1000)
50
50
1500
(5000)
5000
2000
50
50
10
500
(500)
(8000)
7500
500
5000
5000
5000
Sj
00
vo
-------
KJ
VO
O
c Year Current
'o> Country Maximum
tt 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 2018 2019 2020 Level
T3
s I
'si '£
<
ia, and Pacific Islands
East Asia, Austra
Eastern Europe, Caucasus, and Central Asia
Peru (Non-
Urban)
Uruguay
Venezuela
Australia
Cambodia
China (National)
China (Beijing)
China (Hong
Kong)
Fiji
Indonesia
Japan
Malaysia
Nepal
Philippines
Republic of
Korea
Singapore
Thailand
Vietnam
Albania
Belarus
Bosnia &
Herzegovina
Croatia
Russia
Turkey
Armenia
Azerbaijan
Georgia
Kazakhstan
Kyrgyzstan
Serbia
Uzbekistan
7
7
7
7
2000
2000
7
500
7
5000
100
3000
10000
7
500
500
500
10000
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
2000
7
50
7
5000
100
3000
10000
7
500
500
500
10000
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
2000
7
50
7
5000
100
(500)
10000
7
500
500
500
2000
7
7
7
7
7
7
7
7
7
7
7
7
7
7
?
7
500
?
2000
?
50
7
5000
100
(500)
10000
7
500
500
500
2000
?
?
7
7
7
7
?
7
?
7
?
?
7
7
?
7
500
?
2000
?
50
7
5000
100
(500)
10000
7
500
500
500
2000
?
?
7
7
7
7
?
7
?
7
?
?
7
7
7
7
500
7
2000
7
50
7
5000
50
(500)
10000
7
500
500
500
500
7
7
7
7
7
7
7
7
7
7
7
7
7
3000
8000
5000
50
2000
?
50
7
5000
50
(500)
10000
7
500
500
500
500
?
?
7
7
7
7
?
7
?
7
?
?
7
3000
8000
5000
50
2000
7
50
500
j_jOO
50
(500)
10000
7
500
50
500
500
7
7
7
7
7
7
7
7
7
7
7
7
7
3000
8000
5000
50
2000
50
50
jjOO
50
(500)
10000
7
500
50
500
500
7
7
7
7
500
SO/
1000
350
7
7
2000
(350)
7
7
3000
8000
5000
10
2000
50
50
jjOO
50
(500)
10000
7
500
50
500
500
?
?
7
50
350
SO/
1000
50
7
350
2000
(350)
10000
7
1500
50
2000
350
50
50
(500)
50
50
10000
500
50
50
500
500
350
50
350
50
50
SO/
1000
10
2000
50
10
(350)
(350)
5000
50
50
500
150
10
50
2000
10
50
50
10
10
2000
50
2000
50
2000
10
500
5000
10
1500
4000
500
2000
350
350
2000-
5000
10000
400-2000
3
a.
5'
*>,
-------
c Year Current
'o> Country Maximum
tt 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 2018 2019 2020 Level
South Asia
Southwest Asia and North Africa
Sub-Saharan Africa
Afghanistan
Bangladesh
India
India (Selected
Areas)
Nepal
Pakistan
Sri Lanka
Algeria
Bahrain
Egypt
Iran
Iraq
Israel
Jordan
Kuwait
Lebanon
Libya
Morocco
Oman
Palestinian
territories
Qatar
Saudia Arabia
Syria
Tunisia
United Arab
Emirates
Yemen
Angola
Benin
Botswana
Burkina Faso
Burundi
Cameroon
No existing or planned standards.
?
2500
10000
7
10000
7
2500
10000
7
10000
7
2500
10000
7
10000
7
2500
10000
7
3000
7
2500
10000
7
?
7
500
350
10000
7
7
5000
500
350
(500)
7
?
500
350
(500)
7
7
500
350
(500)
7
7
500
350
(500)
7
?
350
50
(500)
7
500
(350)
7
500
No existing or planned standards.
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
?
?
7
?
7
?
?
?
?
7
?
7
?
?
7
7
7
7
7
7
7
?
?
7
?
7
?
?
7
7
7
7
7
7
7
7
7
7
7
7
7
7
?
?
7
?
10
?
(50)
(500)
(5000)
(5000)
(10000)
350
(50)
(50)
(50)
(10)
No existing or planned standards.
7
7
7
7
7
7
7
7
7
7
?
7
7
?
7
7
7
7
7
?
7
7
7
7
7
7
(50)
7
50
(50)
(1000)
50
(50)
No existing or planned standards.
7
7
7
7
7
7
7
7
7
7
7
7
7
7
7
?
7
7
?
?
?
7
7
?
?
7
7
7
7
7
?
7
7
10000
?
7
7
7
350
7
7
7
7
350
7
?
7
(50)
10
(SO)/
500
500
7
(50)
500
7
(50)
10
7
(50)
7
(50)
50
(50)
50
No existing or planned standards.
7
7
7
?
?
7
3000
No existing standards.
3500
3500
3500
3500
3500
3500
3500
3500
3500
50
No existing or planned standards.
No existing standards.
3500
3500
3500
3500
3500
3500
3500
3500
3500
50
No existing or planned standards.
7
7
7
?
?
7
?
7
7
?
5000
900
500
10000
10000
25000
10
(7000-
10000)
2000
(5,005,000)
1500
50
500
(10000)
500
5000
7000
5000
350
10000
(10000)
(500)
(5000)
(5000)
5000
-------
Sj
vo
Sj
c Year Current
'o> Country Maximum
tt 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 2018 2019 2020 Level
Sub-Saharan Africa
Cape Verde
Central African
Republic
Chad
Coted'lvoire
Dem. Rep. of
Congo
Djibouti
Equitorial
Guinea
Eritrea
Ethiopia
Gabon
Gambia
Ghana
Guinea
Guinea Bissau
Kenya
Liberia
Madagascar
Malawi
Mali
Mauritania
Mauritius
Mozambique
Namibia
Niger
Nigeria
Republic of the
Congo
Senegal
Sierra Leone
South Africa
Tanzania
Togo
Uganda
No existing or planned standards.
No existing or planned standards.
No existing standards.
?
?
7
7
7
7
?
7
?
7
7
7
?
7
7
7
7
7
?
7
5000
5000
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
50
50
50
No existing or planned standards.
7
7
7
7
7
7
?
7
?
7
7
7
?
7
7
7
7
7
?
7
7000
(5000)
No existing or planned standards.
No existing or planned standards.
7
7
7
7
7
7
7
7
7
7
5000
No existing standards.
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
50
50
No existing or planned standards.
7
7
7
7
7
7
7
7
(5000)
No existing standards.
7
7
7
7
7
7
7
?
7
?
7
7
7
?
7
7
7
500
7
5000
No existing standards.
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
50
50
No existing or planned standards.
7
7
2500
7
7
7
7
7
7
7
500
No existing or planned standards.
7
7
7
7
7
7
7
7
7
7
7
7
?
7
?
?
?
7
?
?
7
7
7
7
?
7
?
?
7
7
7
7
7
7
7
7
?
7
?
?
(1330)
(1000)
No existing standards.
7
7
7
7
7
7
7
7
7
7
7
7
7
?
7
?
7
?
7
?
7
7
7
7
500
?
7
?
-50
7
7
7
7
7
7
?
7
?
5000
10000
5000
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
3500
50
50
50
50
(3000)
(3000-
5000)
5000
5000
(5000)
5000
(5000-
8000)
7000
10000
(8000)
5000
(5000)
(5000)
10000
(5000)
500
500
(10000)
(5000)
2500
500
(500)
10000
5000
10000
5000
(5000)
500
5000
10000
5000
-------
c
'g> Country
cc
1
Year Current
Maximum
2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 2018 2019 2020 Level
Zambia ?
Zimbabwe
7
?
7
?
7
?
7
?
7
?
7
?
7
?
7
7
500
7
7500
7500
Notes:
1.
2.
3.
4.
5.
Source Material:
1
2
3
Parentheses indicates that fuel of a given sulfur level is available or sold in that country, though the national standard may differ.
Strikethrough numbers indicate that fuel with sulfur in excess of the local standard is commonly sold.
Gray-shaded numbers indicate intergovernmental agreement on future standards.
Underlined numbers refer to agreements made by national ministerial-level officials.
Italicized numbers refer to agreements made by national officials below ministerial level.
Partnership for Clean Fuels and Vehicles, United Nations Environmental Programme (http://www.unep.org/pcfv).
Krylov et al. (2005) Current problems. Low-sulfur diesel fuels: pluses and minuses. Chemistry and Technology of Fuels and Oils 41: 423-428.
Regional Environmental Centre for the Caucasus. (2008) Fuel quality and vehicle emission standards overview for the Azerbaijan Republic, Georgia, the Kyrgyz
Republic, the Republic of Armenia, the Republic of Kazakhstan, the Republic of Moldova, the Republic of Turkmenistan, the Republic of Uzbekistan, and the Russian
Federation, (http://www.unep.org/pcfv).
4. Rojas-Bracho, L. (2009) An international strategy for black carbon controls in the transport sector. The case of Mexico. Presented at International Workshop on
Black Carbon, November 30, 2009. (http://www.theicct.org/information/workshops/international_workshop_on_black_carbon).
5. West and Central African Regional Framework Agreement on Air Pollution, Abidjan, Ivory Coast, July 20-22, 2009.
6. UNEP(2008b). Final Report of the 16th Meeting of the Forum of Ministers of the Environment of Latin America and the Caribbean. UNEP/LAC-IGWG.XVI/9.
n
o
3
O
3
2
S"
n
n
Q
—i
O-
O
3
O
I—
3
•n
vo
Oj
-------
Appendix 4
Table A4-2. International Fuel Sulfur Limits for C3 Marine Fuel, by Target Year. (Source: U.S. EPA)
Global 1 EGA
2004
2012
2020
45,000 ppm
35,000 ppm
5,000 ppm
2005
2010
2015
15,000 ppm
10,000 ppm
1,000 ppm
294
Report to Congress on Black Carbon
-------
Appendix 5
U.S. Emission Standards for Mobile
Sources
Table A5-1. On-Road Regulations for PM Control. (Source: U.S. EPA)
Regulation Sector T^tT Qt ^ , Units
9 Applicable Standard
TierO rule
TierO rule
1985 Heavy Duty Diesel Rule
1991 Heavy Duty Diesel Rule
1991 Heavy Duty Diesel Rule
Tier 1 rule
1994 Heavy Duty Diesel Rule
1994 Heavy Duty Diesel Rule
1994 Heavy Duty Diesel Rule
2001 Heavy Duty Diesel Rule
NLEVrule
NLEVrule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
light duty diesel trucks 1
light duty diesel trucks 2
heavy-duty highway Cl engines
heavy-duty highway Cl engines
urban buses
light duty diesel trucks 1 and 2
heavy-duty highway Cl engines
urban buses
urban buses
heavy duty onroad Cl engines
light duty diesel LEV cars and trucks
light duty diesel ZLEV cars and trucks
LDVLLDT bins 2-6
LDVLLDT-50%bins2-6
LDVLLDT -75% bins 2-6
LDVLLDT -bins 2-6
LDVLLDT -25% bins 7-8
LDVLLDT -50% bins 7-8
LDVLLDT -75% bins 7-8
LDVLLDT -bins 7-8
LDVLLDT -bin 9
LDVLLDT -bin 10
1981-1993
1981-1993
1988-90
1991-1993
1991-1993
1994-1999
1994-2006
1994-1995
1996-2006
2007+
1999-2003
1999-2003
2004
2005
2006
2007
2004
2005
2006
2007
2004-2006
2004-2006
0.26
0.13
0.60
0.25
0.10
0.10
0.10
0.07
0.05
0.01
0.08
0.04
0.01
0.01
0.01
0.01
0.02
0.02
0.02
0.02
0.06
0.08
g/mile
g/mile
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/mile
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/mile
g/mile
g/mile
g/mile
g/mile
g/mile
g/mile
g/mile
g/mile
g/mile
g/mile
g/mile
Report to Congress on Black Carbon
295
-------
Appendix 5
Table A5-2. On-Road Gasoline-Vehicle Regulations. (Source: U.S. EPA)
Regulation Sector T^tT Qt ^ , Units
9 Applicable Standard
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2000 Tier 2 Rule
2001 Heavy Duty Diesel Rule
2001 Heavy Duty Diesel Rule
Onroad Gasoline - 25% bins 2-6
Onroad Gasoline - 50% bins 2-6
Onroad Gasoline - 75% bins 2-6
Onroad Gasoline - bins 2-6
Onroad Gasoline - 25% bins 7-8
Onroad Gasoline - 50% bins 7-8
Onroad Gasoline - 75% bins 7-8
Onroad Gasoline - bins 7-8
Onroad Gasoline - bin 9
Onroad Gasoline- bin 10
Heavy Duty Onroad Gasoline 50%
Heavy Duty Onroad Gasoline 100%
2004
2005
2006
2007
2004
2005
2006
2007
2004-2006
2004-2006
2008
2009
0.01
0.01
0.01
0.01
0.02
0.02
0.02
0.02
0.06
0.08
0.01
0.01
g/mile
g/mile
g/mile
g/mile
g/mile
g/mile
g/mile
g/mile
g/mile
g/mile
g/bhp-hr
g/bhp-hr
Table A5-3. Nonroad Diesel Regulations. (Source: U.S. EPA)
Regulation Sector T^tT Qt PH A Units
9 Applicable Standard
2004 Nonroad Tier 4
2004 Nonroad Tier 4
2004 Nonroad Tier 4
2004 Nonroad Tier 4
2004 Nonroad Tier 4
2004 Nonroad Tier 4
2004 Nonroad Tier 4
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
1998 Nonroad Diesel Engines Rule
Nonroad Diesel, hp<25
Nonroad Diesel, 25560
Nonroad Diesel, kW>560
2008
2008
2012
2011
2011
2015
2015
2000
2005
2000
2005
1999
2004
2004
2003
1996
2003
1996
2001
1996
2002
2000
2006
0.3
0.22
0.01
0.01
0.075
0.02
0.03
1
0.8
0.8
0.8
0.8
0.6
0.4
0.3
0.54
0.2
0.54
0.2
0.54
0.2
0.54
0.2
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
g/kW-hr
296
Report to Congress on Black Carbon
-------
U.S. Emission Standards for Mobile Sources
Table A5-4. Locomotive Regulations.3 (Source: U.S. EPA)
Regulation Sector T^tT Qt ^ , Units
9 Applicable Standard
2008 LocoMarine Rule
2008 LocoMarine Rule
2008 LocoMarine Rule
2008 LocoMarine Rule
2008 LocoMarine Rule
2008 LocoMarine Rule
2008 LocoMarine Rule
2008 LocoMarine Rule
2008 LocoMarine Rule
2008 LocoMarine Rule
1997 Locomotive Stds
1997 Locomotive Stds
1997 Locomotive Stds
1997 Locomotive Stds
1997 Locomotive Stds
1997 Locomotive Stds
Locomotive - Line Haul
Locomotive- Switch
Locomotive - Line Haul
Locomotive- Switch
Locomotive - Line Haul
Locomotive- Switch
Locomotive - Line Haul
Locomotive- Switch
Locomotive - Line Haul
Locomotive- Switch
Locomotive - Line Haul
Locomotive- Switch
Locomotive - Line Haul
Locomotive- Switch
Locomotive - Line Haul
Locomotive- Switch
Tier 0(1 973-1 992)
Tier 0(1 973-2001)
Tierl (1993-2004)
Tier 1 (2002-2004)
Tier 2 (2005-2011)
Tier 2 (2005-2010)
Tier 3 (2012-2014)
Tier 3 (2011-2014)
Tier 4 (201 5+)
Tier 4 (201 5+)
Tier 0(1 973-2001)
Tier 0(1 973-2001)
Tier 1 (2002-2004)
Tier 1 (2002-2004)
Tier 2 (2005+)
Tier 2 (2005+)
0.22
0.26
0.22
0.26
0.10
0.13
0.10
0.10
0.03
0.03
0.60
0.72
0.45
0.54
0.20
0.24
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
g/bhp-hr
i Note that in the 2008 Locomotive/Marine Rule, EPA revised its emission standards for remanufactured locomotives to redesignate as
Tier 1 most locomotives originally built between 1993 and 2001. EPA determined that these locomotives could be readily retrofitted to
meet the more stringent Tier 1 standards when remanufactured.
Report to Congress on Black Carbon
297
-------
Appendix 5
Table A5-5. Commercial Marine Regulations. (Source: U.S. EPA)
n . .. _ . Model Years PM ....
Re9ulatlon Sector Applicable standard Umts
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 3
2008 LocoMarine Rule - Tier 4
2008 LocoMarine Rule - Tier 4
2008 LocoMarine Rule - Tier 4
1999 C1&C2 Marine Engine Rule
1999 C1&C2 Marine Engine Rule
1999 C1&C2 Marine Engine Rule
1999 C1&C2 Marine Engine Rule
1999 C1&C2 Marine Engine Rule
1999 C1&C2 Marine Engine Rule
C1 Commercial Stnd Power Density, <19kW Max power,
<0.9L/cylinder
C1 Commercial Stnd Power Density, 19<75kW Max
power, <0.9L/cylinder
C1 Commercial Stnd Power Density, 75<3700kW Max
power, <0.9L/cylinder
C1 Commercial Stnd Power Density, 75<3700kW Max
power, 0.9<1.2L/cylinder
C1 Commercial Stnd Power Density, 75<3700kW Max
power, 1.2<2.5L/cylinder
C1 Commercial Stnd Power Density, 75<3700kW Max
power, 2.5<3.5L/cylinder
C1 Commercial Stnd Power Density, 75<3700kW Max
power, 3.5<7.0L/cylinder
C1 Rec and Com High Power Density, <19kW Max power,
<0.9L/cylinder
C1 Rec and Com High Power Density, 19<75kW Max
power, <0.9L/cylinder
C1 Rec and Com High Power Density, 75<3700kW Max
power, <0.9L/cylinder
C1 Rec and Com High Power Density, 75<3700kW Max
power, 0.9<1.2L/cylinder
C1 Rec and Com High Power Density, 75<3700kW Max
power, 1.2<2.5L/cylinder
C1 Rec and Com High Power Density, 75<3700kW Max
power, 2.5<3.5L/cylinder
C1 Rec and Com High Power Density, 75<3700kW Max
power, 3.5<7.0L/cylinder
C2,<3700kW,7-<15L/cylinder
C2,<3700kW,15-<30L/cylinder
C1&C2,>3700kW
C1&C2,>3700kW
C1&C2,600to<3700
C1, power <37 kW disp. <0.9
C1,0.9
-------
Appendix 6
International Emission Standards
for Heavy-Duty Vehicles
Heavy-duty on-road diesel vehicles represent
the predominant mobile source of BC in most
areas although nonroad diesel, locomotives and
commercial marine can also be significant. The
following discussion addresses emission standards in
other parts of the world.
Outside the United States, Europe, and Japan, other
nations adopt heavy-duty engine emission standards
developed by these governments using schedules
determined by legislative or executive standards. As
noted earlier, Canada generally adopts U.S. standards
on a timeframe similar to the United States, Australia
also bases its national standards on those developed
in the United States, Europe, or Japan. Outside these
nations, other countries adopt emission standards,
generally based on European standards, albeit on
a different time frame. As discussed in Appendix 4,
countries must ensure that fuel quality is requisite
to allow emissions-reduction technologies to be
implemented.
A number of countries have adopted schedules
for phasing in PM emission standards for heavy-
duty diesel engines that are likely to require
advanced aftertreatment, such as a DPF, to meet
the relevant national standard. In the Americas,
Brazil's PROCONVE P7 standards beginning in the
2012 model year are likely to require advanced
aftertreatment. Russia has adopted standards based
on EURO IV starting in the 2010 model year and
standards based on EURO V in the 2014 model
year. In the Beijing area, China adopted standards
equivalent EURO IV in 2008, and has proposed
adoption of EURO V-equivalent standards in 2012.
In addition, several countries that have applied
for membership in the European Union will adopt
EURO standards if accepted. These countries include
Croatia, Iceland, Macedonia, and Turkey. Other
potential candidate countries that have not formally
petitioned for EU membership include Albania,
Bosnia and Herzegovina, Kosovo, Montenegro, and
Serbia.
Numerous other countries have adopted or
proposed heavy-duty engine emission standards
equivalent to earlier U.S. or EURO emission
standards. In the Americas, these countries include
Argentina, Brazil, Chile, Mexico, and Peru. In the
western Pacific and Asia, these countries include
China, India, the Republic of Korea, Singapore, and
Thailand. In Europe outside of the European Union,
Russia and Turkey have adopted earlier EURO
standards. These countries are making progress in
reducing BC emissions from heavy-duty vehicles.
Figures A6-1, A6-2, and A6-3 show a graph of how
PM emission standards are changing over time
in the Americas, Asia and Australia, and Europe,
respectively.1 The figures also include trend lines,
indicating the general trend of emission standards
over time. As illustrated, most countries with
emission standards in place have introduced
progressively more stringent standards over time.
The scatter around the trend line of each country
reflects differences in standards based on vehicle
type (truck vs. bus), test procedure (e.g., operating
cycle), and/or location (e.g., urban vs. rural).
Beyond nations that have regulations with emission
standards, other nations have been addressing
vehicle emissions in some manner. Some other
nations are adopting emission standards for light-
duty vehicles, generally based on EURO standards.
Others have eliminated or are scheduled to
eliminate lead from gasoline, which enables the
implementation of standards to reduce tailpipe
emissions using catalytic aftertreatment. An example
of this progress is found in Africa, where all nations
have eliminated lead in gasoline. Others have
banned the import of light-duty vehicles without
a catalytic converter or established opacity testing
requirements for cars, trucks, or scooters. This
progress suggests room for additional technology-
based approaches to reducing BC emissions.
Many other countries lack any emission standards.
The reasons for their lack of emission standards
may be attributable to several causes, including
insufficient governmental capacity, poverty and
other economic factors, and government policy.
Many such countries face many other problems
related to economic development, public health,
violence, and authoritarian rule. Addressing BC from
motor vehicle emissions in these locations may
requires attention to factors other than technology.
1 See http://www.dieselnet.com/standards for more information.
Report to Congress on Black Carbon
299
-------
Appendix 6
•5 0.1
C
5 0.01
0.
* Argentina
• Brazil
Chile (National)
x Chile (Partial)
• Mexico
- Peru
United States
Expon. (Argentina)
Expon. (Brazil)
Expon. (Chile (Partial))
Expon. (Mexico)
Expon. (Peru)
Expon. (United States)
0.001
1985
1990
1995
2000
Model Year
2005
2010
2015
Figure A6-1. Heavy-Duty Highway Diesel PM Emissions Standards in the Americas and the Caribbean
(Logarithmic Scale). (Source: U.S. EPA)
0.1
£
o
i.
0.01
0.001
t. • »-,;-,.
•v
1985
1990
1995 2000 2005
Model Year
2010
2015
• Australia
• China
Japan
Republic of Korea
* Singapore
• Thailand
-t- India
Expon. (Australia)
Expon. (China)
Expon. (Japan)
Expon. (Republic of Korea)
Expon. (Singapore)
Expon. (Singapore)
Expon, (Thailand)
Expon. (India)
Figure A6-2. Heavy-Duty Highway Diesel PM Emissions Standards in Asia and Australia (Logarithmic
Scale). (Source: U.S. EPA)
300
Report to Congress on Black Carbon
-------
International Emission Standards for Heavy-Duty Vehicles
0.1
:
a.
I
0.01
0.001
Russia
Turkey
European Union
- Expon. (Russia)
Expon. (European Union)
1985
1990
1995
2000
Model Year
2005
2010
2015
Figure A6-3. Heavy-Duty Highway Diesel PM Emissions Standards in Europe (Logarithmic Scale).
(Source: U.S. EPA)
Report to Congress on Black Carbon
301
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Appendix 7
Research Needs
The chapters of this report have laid out much of
what is known about BC. While there is a great
body of knowledge about BC, its emissions, its
atmospheric properties, its effects on climate, and
its impacts on health, more information on these
topics would lead to enhanced capability to design
policies and control strategies that would be most
beneficial for the environment and for human
health. Many of these research priorities have been
highlighted in this report; these research needs are
summarized here, with the relevant section(s) of
the report calling for each research topic given in
parentheses. These descriptions also give more detail
on the uncertainties and research needs discussed in
sections 12.5 and 12.6.
A7.1 Basic Atmospheric Chemistry
of BC, LAC, and Measurement
Approaches
Establishing standardized definitions of terms
related to carbonaceous aerosols and harmonizing
measurement approaches will reduce the uncertainty
in emissions inventories and result in improved and
more consistent air quality and climate estimates.
Standardized definitions will also encourage the
use of preferred monitoring approaches and more
consistent, less uncertain data, resulting in better
climate and health assessments; currently, a lack
of consistency in connections between BC and
health endpoints could be caused by uncertain or
inconsistently produced data. These definitions can
provide the basis for future reference and equivalent
measurement methods for a regulatory program if
ever needed.
• It is important to develop standardized definitions
of BC, BrC, non-light-absorbing OC and other
light-absorbing aerosols (2.3).
• It is also important to establish corresponding
reference materials for needed particle properties
(2.3, 5.2).
• Interpretation and harmonization of existing
ambient and emissions measurements are
necessary to conform to the specified definitions
and properties to meet the needs of climate and
health assessments (5.2).
While BC absorbs light across the entire solar
spectrum, BrC absorbs primarily in the ultraviolet
range. The quantity of solar energy absorbed
depends upon the molecular structure and the
mass of the relevant BrC compounds present in a
combustion plume or aerosol mixture. OC is usually
the most significant co-emitted pollutant, by mass,
among the major BC emitting sources.
• Estimates are needed to separately characterize
BrC and the non-light-absorbing (scattering)
portions of OC. Characterizations of the
particulate, semi-volatile and perhaps some
volatile ("intermediate VOC") components of non-
light-absorbing (scattering) portions of OC are
especially important to improve inventories, to
improve model estimates of SOA, and to evaluate
models.
• It is necessary to develop and/or improve
instrumentation and measurement techniques to
identify and quantify the physical and chemical
properties (optical, size, number, composition,
and mass) of BrC and LAC, and through
modeling, to estimate the radiative forcing of BrC.
• It is also important to develop the ability to track
the radiative forcing and cloud droplet-forming
potential of particles, from various emissions
sources, as they age in the atmosphere (5.2, 2.6).
There is also a need to improve instrumentation and
measurement techniques to identify and quantify
the properties of BC (optical, size, number, and
mass) taking into consideration different emissions
sources, combustion conditions and varying particle
mixtures (5.2, 4.3, 4.4). Further study is needed
to understand the differences among existing
thermal and optical monitoring and measurement
methods for BC and how they are affected by (a) the
emissions sources and resulting aerosol mixtures
(considering variations in the combustion process
and different fuels for source measurements),
(b) atmospheric processing (aging), and (c) the
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manner in which monitoring is performed. Preferred
protocols for future measurements and monitoring
are desired to ensure more consistent and useful
data collection. Improved source measurement
techniques will facilitate the opportunity for more
measurements, better emission factors and then
better inventories. Multiple methods may be needed
to serve different data needs.
A7.2 Ambient and Emissions Source
Measurement
Emissions source measurements are much more
limited than ambient measurements. Additional
representative source measurements are needed
to better characterize BC emissions (and speciated
profiles) by emissions source, fuel type and
combustion conditions. These data are needed to
improve BC emission factors and inventories, to
create BrC emission factors and inventories, and to
help develop emissions and modeling uncertainties.
Research can also help guide a new focused data
collection effort.
• Methods for using source measurements to
develop emission factors for light-absorbing
carbon (including BrC) - particularly for the major
source categories like open biomass combustion
and mobile sources - need to be explored (5.2,
4.3, 4.4).
Regular speciated PM measurements, including EC
(or BC) measurements, currently occur at very coarse
spatial and temporal scales. More spatially dense
ambient monitoring can help identify unaccounted
for emissions and sources. More temporal estimates
(within and among years) as well as semi-continuous
data will help account for the impacts of emissions
trends. Understanding the effect of historical
emissions changes will be useful in estimating the
need for future emissions reductions to meet various
policy goals.
• More ground-level ambient measurements are
needed, with improved spatial and temporal
coverage, particularly outside the United States.
These data would be useful for a multitude
of reasons, including better understanding of
pollution trends and better linkages between
PM components, such as BC, and health, climate,
ecosystem, and visibility outcomes (5.3, 5.4, and
5.7).
A consolidated global database of BC and OC
ambient and source measurements would be useful
for facilitating the analysis of all existing data and
for ensuring consistent development of inventories
and evaluation of global climate models. A protocol
should be established to improve the quality and
to guide the reporting of meta-data available both
for ambient measurements and source profiles (e.g.
identifying conversions of absorption measurements
to mass and adjustments for sampling artifacts
and or material balance). Consistent reporting of
measurements is desired because the variability
among existing measurement approaches and data
contribute to the uncertainty of domestic and global
emissions estimates and impacts.
• Mechanisms for archiving, consolidating,
and sharing existing measured data on light
absorption and scattering of particles, available
globally, are needed to reduce the need for new
measurements and to produce better air quality
and climate characterizations (5.2, 5.7).
Full chemical composition emissions measurements
for the important carbonaceous combustion
sources, including the identification of carbonaceous
co-pollutants in addition to BC (OC, BrC)
measurements would promote better understanding
the atmospheric aging and climate-relevant
properties of carbonaceous aerosols as well as
would be essential in constructing the emissions
inventories needed for assessment of any co-
benefits and tradeoffs from BC mitigation efforts
(5.2, 4.3, 4.4, 6.4).
A7.3 Emissions Inventories
A7.3.1 Global
Given the importance of emissions inventories
in understanding the impacts of BC and control
measures, more accurate representations of BC
emissions globally is critical for more accurate
estimates of BC's impacts. Sources of BC outside
the United States are responsible for 94% or so
of the current emissions globally and this number
is expected to increase in the future. The work
of Bond et al. (i.e., 2004) is widely recognized as
the "best currently available" global BC inventory.
That work is based on combinations of fuel,
combustion type, and emissions controls and their
prevalence on a regional basis. The 2004 work
(which represented 1996 fuel-use data) has since
been updated to reflect more recent years and to
improve the emission factors and usage patterns
for some sources. However, there remain significant
uncertainties for BC inventories in developing
countries and globally. There are several ways in
which these inventories can be improved. Suggested
areas for further research largely focus on improving
these emissions by gathering more data both on the
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emissions side and on the "usage" (activity level) side
as well on determining how best to allocate these
emissions both spatially and temporally:
• For mobile sources, there is a need to better
characterize the on-road and nonroad fleet in
urban and rural areas for different countries
including commercial marine and locomotives.
• Emissions estimates for nonroad applications
need to be developed and collated separately
from on-road operations (4.4.1).
• In the case of biomass burning, all fires need to be
better captured for their emissions characteristics
and activity level.
• A careful review of usage patterns needs to be
conducted to ensure that appropriate "activity"
levels are applied to emission factors to arrive at
final emissions estimates (4.4.3).
• Small(er) sources are especially poorly
represented on a regional basis, and better
characterization of emissions from residential
cookstoves, in-use mobile sources, small fires,
smaller industrial sources such as brick kilns, and
flaring emissions is needed.
- Many of these "small" sources could have
relatively high BC emission factors (4.4.1 and
4.4.3).
- For sources such as cookstoves, improved
characterization depends critically on field-
based measurements of emissions from in-use
sources.
- In addition, usage patterns need to be
reviewed to ensure that appropriate "activity"
levels are applied to emission factors to arrive
at final emissions estimates.
- Finally, fuller incorporation of regional
inventories into global inventories could
improve country- and region- specific
emissions estimates.
• Steps need to be taken to better engage
international scientists, governments, and
regulatory agencies in collaborations of
technology assessments to help improve base-
year and out-year global emissions.
- Merging regional inventories with global
estimates should be explored in an attempt
to improve country- and region-specific
emissions estimates in global inventories
(4.4.3 and 4.4.4).
- Initiating an international forum in which
scientists and governments can more readily
and routinely engage to help facilitate
and share this research would expedite
emissions inventory improvement and help to
harmonize estimation methods across world
regions.
A7.3.1.1 Domestic
While domestic emissions of BC and PM are
generally better characterized than global emissions,
considerable uncertainty remains for these
estimates and there are several aspects of domestic
inventories that need improvement.
• More information on both emission factors and
usage would be helpful. In particular, emissions
from key industrial sources, flaring, residential
heating, and open biomass burning remain
poorly characterized.
• In general, mobile source emissions are among
the best characterized (especially in developed
countries), but improved information is still
needed for some sectors, most likely through
increased testing and data acquisition. (These
needs are all discussed in Section 4.3.2.)
- Special attention should be given to PM
mass and composition from nonroad sources
(gasoline and diesel), aircraft (including
in-flight emissions), commercial marine
especially C3 (ocean going), and locomotives
including both current and, where available,
future technologies.
- Characterization of emissions from newer-
technology on-road diesel and gasoline
vehicles would also be useful, as would better
emissions characterization (including fleet
fraction) of high-emitting vehicles/engines
(so-called "super-emitters").
- Finally, the improved characterization of
emissions at low ambient temperature and
with different fuels (including renewables)
would improve the understanding of present
and future emissions from mobile sources.
These data can then be used to develop newer and
more accurate emission factors for these key areas
within mobile sources. These new data could also
enable construction of emissions models specifically
for EC for nonroad sources similar to what exists for
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on-road (and rely less on using PM2.5 models and
translating to EC via a separate database in which
speciated emissions from sources are archived).
A7.3.2 Uncertainty Analysis
For both global and domestic EC/BC emissions,
analyses of uncertainty would aid in determining
which sectors have the strongest estimates of
BC emissions, which in turn could lead to easier
decisions on mitigation options. Bond et al. (2004)
have done a Monte Carlo-type uncertainty analysis
for global emissions and have estimated that
uncertainty in BC emissions inventories is generally
on the order of a "factor of 2," and more work
along these lines is needed. A starting point on the
domestic side would be to qualitatively rank sources
by considering the strengths and weaknesses of how
the BC emissions were assembled. This could lead to
a research task of doing a more rigorous uncertainty
analysis on the global estimates. This method
could also be applied to future-year projections of
emissions.
• Quantitative measures for describing uncertainties
in emissions should be advanced, as was
suggested in Section 4.4.4.
• As a first step, a qualitative description of
uncertainty in emissions estimates, sector-
by-sector, for domestic emissions inventories
including models should be performed.
• The uncertainty that stems from combining BC
data sets collected by several measurement
techniques (for regional and global inventories)
should be estimated.
• Finally, the Monte Carlo-type work by Bond
et al. (2004) should be extended to determine
uncertainties in global emissions estimates.
A7.4 Ambient Observations, Including
Deposition
While many kinds of additional measurements
would improve our understanding of BC in the
atmosphere, a few specific types of measurements
were highlighted in this report as able to fill some
important gaps in the current understanding of
BC. BC's vertical distribution (and its impact on
surface and atmospheric radiative forcing and cloud
formation) is one of the important uncertainties
in assessing BC's overall impact. Similarly, the
deposition of BC is a source of uncertainty in
determining BC's overall impacts. For example, a
recent NOAA/GFDL modeling study (Liu et al., 2011)
showed that the simulated vertical column of BC
concentrations over the Arctic is highly sensitive to
different parameterizations of deposition properties.
• Better characterizing the vertical distribution
of BC would allow for a fuller understanding
of its climatic impacts, which are dependent
on its vertical distribution, and would further
understanding of the discrepancies between
models and observations (5.5 and 5.7).
• Research to inform the characterization of BC wet
scavenging and dry deposition rates, deposition
on snow and ice and resulting radiative forcing,
albedo, and hydrological changes (2.6, 5.6.2,
5.6.5) would improve the modeling of these
important processes.
A7.5 Modeling
Some aspects of the modeling of BC in the
atmosphere, and its effects on climate, are better
understood than others. BC's aging/mixing states
(internal, external, or core-shell) and its indirect and
semi-direct effects are likely the largest uncertainty
in assessing BC's RF and climate impacts.
• Improving the modeling of BC's direct effects due
to aging/mixing states assumptions, semi-direct
effects (on vertical mixing, clouds, and differential
heating at surface and the atmosphere), and
indirect effects (on cloud formation, lifetime,
albedo, etc.) would be very useful for improving
representations of BC's environmental impacts
(2.6).
BrC is also a heretofore under-studied aspect of
the impacts of carbonaceous aerosols. Inclusion of
BrC in climate models, and analysis of implications
for net forcing, including a sensitivity analysis of
the upper and lower reasonable bounds for BrC
absorption, would enable a fuller accounting of the
climatic impacts of carbonaceous aerosol sources.
• Reporting column data by wavelength may aid
model-observation comparison as BC and BrC
differ in terms of peak absorption (2.3, 5.6.2).
• Coupled with experimental estimates of BrC
emissions and laboratory estimates of BrC
scattering and absorption, this should clarify the
magnitude of the cooling offset due to organic
carbon co-emissions, which is important for
determining net forcing of abatement measures.
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A7.6 Climate Impacts
Further research on the impacts of BC and
carbonaceous aerosols on climate, with an eye
toward filling in the continuum from source to
impact, would enable policies that are more likely
to have beneficial impacts. Several aspects of this
issue were highlighted in the chapters of this report.
Emissions of other pollutants from the same sources
can lead to difficulties in determining whether a
control strategy will result in radiative cooling or
warming. In addition, there are indications that
near the Arctic and other snow-covered regions,
the net effect of any mitigation measure including
measurable amounts of BC is much more likely to be
net cooling.
• Better characterization of the sources as well as
the "total" radiative effect of the control measure
will reduce the possibility of unintentional
warming as well as lead to more efficient policy.
• Research is needed to characterize the range of
possible control strategies and evaluating the
probability that a given measure will result in net
cooling (6.4). Specific focus should be placed on
the location (especially latitude) of the proposed
change in emissions, especially for near-Arctic
emissions. Specific measures would be the goal,
but sectorial-level analysis would also be useful.
One important aspect of BC's impacts on climate is
its role in snow and ice melting, this is of particular
relevance in areas where BC deposition may
affect snow pack that influences the availability of
water resources for downstream populations (e.g.,
California, Himalayas and Tibetan Plateau, Ancles,
high African mountains) as well as in the Arctic.
Focused research on the role of deposited BC
could shed light on the effectiveness of mitigation
measures for protecting water resources and snow in
sensitive regions.
Specific contributors to the overall uncertainty in
BC's impacts include aerosol mixing state and cloud
impacts. The biggest uncertainty in direct effect
calculations is due to aging/mixing of particles,
and the biggest uncertainty overall may be cloud
interactions. Continued research is needed on
the radiative properties of BC and co-emissions,
especially regarding cloud interactions and effects of
aging/mixing (2.6).
Non-radiative impacts of BC are more poorly
understood than its radiative impacts; even if the net
radiative effect of a given measure is near-zero, there
may be other climatic impacts. Continued research
is needed on the non-radiative effects of BC and
other aerosols, especially regarding precipitation/
hydrological interactions and dimming (2.6.3).
A7.7 Metrics
It is difficult to apply climate metrics developed
for GHGs to BC and other short-lived forcers
as many of the fundamental assumptions that
go into the calculation of these policy-re levant
metrics for long-lived GHGs are not appropriate for
application to BC. Though "alternative" metrics have
been proposed for BC, none is yet widely utilized.
Appropriately tailored metrics for BC are needed in
order to quantify and communicate BC's impacts
and properly characterize the costs and benefits of
BC mitigation. Improved metrics could incorporate
non-radiative impacts of BC, such as impacts
on precipitation. Similarly, given BC's (and other
aerosols') direct impacts on human health, health
outcomes could also be incorporated into such a
metric. Developing methods to quantify the benefits
of BC mitigation on both climate and health would
encourage policy decisions that factor in climate
and health considerations simultaneously, within a
unified framework.
Ways to quantify the various impacts of BC into a
unified framework or to compare BC to other SLCFs
and GHGs would enable quantification of how BC
mitigation leads to the attainment of various policy
goals:
• Explore the use of emissions source
measurements and/or emissions inventory
estimates to determine whether (or how)
measures like OC/EC ratios, BC (as measured
currently), and other LAC forms can be utilized
to inform development of metrics to prioritize
emissions sources for mitigating the climate
effects of PM emissions (5.2, 2.7.3). Simple
emission-related measures are useful screening
tools that permit mitigation decisions without the
need to run complex climate models. Currently,
such measures are used but with unknown, but
probably large, uncertainty. For example, using
directly emitted OC as an indicator of scattering
from aerosols does not acknowledge the role of
SOA or OC's BrC component.
• There is a need for an analysis of the implications
of using different metric choices on future
emissions and climate, examining the benefits
of near-term versus long-term temperature
abatement and therefore short-lived vs. long-
lived gas abatement (2.7.3,2.7.4). One example
of this would be a system where an economic
model is run with different GWP/GTP values
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for BC, and then the emissions are fed into a
climate model. Key issues involve the relationship
between different metrics values and economic
impacts, and in terms of the temporal pattern of
temperature implications.
• Given the differences in the mechanisms by which
BC and GHGs impact climate and human health,
there is a need for improved quantification and
valuation of specific climate change impacts that
could be attributed to BC emissions, including
analysis of how these differ from quantification of
GHG reduction benefits.
• Additionally, development of approaches to
identify and properly account for co-benefits
in mitigation benefit analyses would more
accurately capture the climatic impacts of policy
options (2.7.3, 2.7.4). Key differences between
BC and other GHGs for valuation purposes
include the regional specificity of BC impacts,
the differences in vertical distribution of forcing,
and cloud interactions. Human health effects
and precipitation, visibility, and dimming
effects are also cited as benefits of BC control.
Including these benefits could improve metrics
development and would move these metrics
beyond simply climatic endpoints to a whole host
of environmental and health goals.
A7.8 Health
A great deal of research on the health impact
of PM2.5 and specific PM components has been
conducted over the past 15 years, and these topics
have already been identified as priorities by EPA
in the context of its periodic reviews of the U.S.
national ambient air quality standards for PM. While
the scientific record is robust in many respects,
there are still important unanswered questions
about the relative toxicity of different constituents.
Also, research continues to inform the overall
understanding of the magnitude and nature of PM
health impacts, including more precise quantitative
information about the relationship between indoor
and ambient concentrations and health impacts.
Continued investment in this research is important,
and there is a particular need for more studies in
developing countries.
In addition to the ongoing research on the health
impacts of PM components such as BC, work is
needed on linking the many types of impacts from
a particular emissions sector. Research is needed to
quantify the integrated climate and health impacts
of individual economic sectors and domestic and
international mitigation measures, accounting
for the full mixture of emissions, both direct and
indirect climate effects, and both indoor and
outdoor exposure (2.7.3.5, 2.7.4, 3.3, and 3.4). This
includes the complex mix of organic compounds
that comprise the co-emitted species from BC
sources such as fossil fuel and biomass combustion.
This would help elucidate where the greatest
opportunities to benefit human health and the
environment lie, with respect to BC mitigation.
A7.9 Mitigation Technologies and
Measurements
Of key importance is research on which BC
mitigation strategies are most cost-effective
and beneficial for public health and climate.
The necessary continued research on mitigation
includes more information on costs and benefits
of mitigation by sector, and development of new
or improved mitigation technologies for various
sectors. This research would in turn allow for
comparisons across sectors in terms of costs
and benefits for climate and human health. The
following research needs are for specific source
sectors.
A7.9.1 Stationary
While emissions from controlled domestic sources
are relative well understood, there are ways in which
estimates could be improved. Additional source
testing and development of improved emission
factors (9.8) would result in higher quality emissions
inventories for stationary sources in the developed
and the developing world. This would improve our
estimates of the effectiveness of traditional control
equipment (baghouses, electrostatic precipitators)
in reducing the BC fraction of PM. Along with
improved activity level estimates, it would improve
overall emissions inventory estimates for the
industrial sector. Key categories would include
coke production, brick kilns, and oil and gas flaring.
Profiles for stationary sources in the developing
world are especially uncertain.
A7.9.2 Mobile
Mobile sources, especially on an international
basis, are an important source of EC emissions
and improved, more effective, and more cost-
effective control technology would result in wider
adoption and likely benefits for both human health
and climate change mitigation. It is important to
continue the development of current mobile source
control technology (such as diesel particulate filters
for in-use vehicles) targeted specifically for EC,
including the ability to assure adequate durability
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and lower costs. Also, the applicability to more
existing engines, especially nonroad, locomotive,
and C1/C2 marine, is needed. Control technology for
C3 marine is also needed. The speciated emissions
profiles that result from the use of existing control
technologies are currently not well characterized
(8.6).
A7.9.3 Cookstoves
Given the ubiquity of high-emitting cookstoves in
the developing world, it is not surprising that there
are many opportunities for research into mitigating
cookstove emissions. First, linking cookstoves to
climate outcomes requires a clear understanding
of the emissions from particular stoves and fuels.
Currently, such information is lacking, and most of
the testing that is occurring is in laboratory settings.
Therefore, there is a need for laboratory and field
testing to characterize emissions (BC, OC, CH4, other
constituents) from different types of cookstoves,
based on stove design, fuel type, and usage patterns
(10.4.1,10.4.2). Expanded lab and in-field testing
data are needed to clarify what constitutes a "clean"
stove/fuel.
Because of the difference in fuels among regions, as
well as the difference in sensitivity of the local and
regional environment to climate forcers, there is a
need for regional-level studies of the net climate
impacts of cookstove emissions from different
cooking stove-fuel combinations, including linkages
to radiative forcing, glacial melt, and precipitation
impacts (10.4.1,10.4.2). Studies evaluating the
extent to which emissions from stoves are linked to
climate impacts at the local/regional level would help
clarify which cookstove mitigation efforts would be
beneficial for climate, and what technologies and
fuels would be needed to achieve maximum climate
benefits.
In order to make good policy decisions about
preferred interventions and investments, there is
need to better understand the linkages between
emissions changes and health benefits. Improved
dose/response information would enable
policymakers to target specific improved stoves
and fuels for development and dissemination.
Specifically, research is recommended on examining
dose-response relationships between emissions of
various cookstove emission constituents and health
endpoints of concern (ALRI/pneumonia, COPD,
cardiovascular disease, cancer, etc.) (10.4.1,10.4.2).
A7.9.4 Residential Heating
There are some uncertainties associated with the
emissions from residential heating. More research
on the composition of particles from residential
heating (10.3.1) would enable better quantification
of the benefits of mitigation from the residential
sector. More data on the optical properties of these
aerosols would be especially useful. Related to this
uncertainty in emissions composition is the effect
that different heating technologies and abatement
options have on the chemical composition of
carbonaceous aerosol emissions (10.3.1). There
is currently little data on whether the improved
stoves used for air quality purposes equally reduce
all PM components, or whether some are reduced
preferentially. These changes in composition can
be measured using existing (or new) techniques for
emissions speciation.
Similarly, the performance of residential heating
appliances as they age has not been well
documented. Given that the lifetime of wood stoves
is on the order of several decades, more complete
long-term performance data on the emissions
from older heating appliances would enable more
accurate emissions and impact projections into the
future (10.3.2).
A7.9.5 Biomass Burning
Globally, a major fraction of BC emissions come
from biomass burning, and yet biomass burning
emissions are especially uncertain. In order to
understand the efficacy of mitigation options,
biomass burning emissions should be better
characterized. Specifically, there is a need for
additional measurements and biomass burning
emissions and activity factors as a function of size
and duration of the fire, fuel type, fuel conditions,
fire phase, and meteorological conditions on the day
of the burn and other significant variables (11.3).
There is a need for more information about total
area burned in each fire category in the United
States and globally, and also a need for additional
fire activity data on a broad scale. A complete
impacts assessment of biomass combustion
emissions requires inventories that include both BC
and BrC, along with co-pollutants that may offset
warming, and other materials that are implicated
in human health and ecosystems impacts. More
information on the plume rise of such fires would
also improve their representation in chemical
transport models.
Regarding actual mitigation measures for
biomass burning, there is a need for an analysis
of the efficacy of existing and proposed methods,
including analysis of total life-cycle impacts. An
assessment of the efficacy, and any unintended
consequences due to the implementation, of
proposed measures for biomass smoke mitigation,
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Appendix 7
beginning with a synthesis of available research combination with others, for ecosystems impacts,
results (11.5.1,11.6) would be especially useful in (2) total life-cycle studies of the net climate forcing
considering mitigation options. This type of analysis impacts arising from the use of these methods, or
could conceivably include (1) total life-cycle studies (3) total life-cycle studies of the economic impacts
of mitigation methods, both individually and in of the use of these methods.
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SEPA
United States
Environmental
Protection Agency
Office of Airand Radiation
Office of Air Quality Planning and Standards
Research Triangle Park, NC
Publication No. EPA-450/R-12-001
March 2012
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