Report to Congress on Black Carbon

March 2012

Department of the Interior, Environment, and Related Agencies 

Appropriations Act, 2010

United States 

Environmental 

Protection Agency
Printed on 100% recycled/recyclable process chlorine-free paper with 100% post-consumer fiber using vegetable oil-based ink.
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

EPA-450/R-12-001

March 2012

Report to Congress on Black Carbon

Report to Congress on Black Carbon i

Acknowledgments

Venkatesh Rao

Joann Rice

Marcus Sarofim

Joseph Somers

Charlene Spells

Sara Terry

Matthew Witosky

Lead Authors

Erika Sasser (Chair)

James Hemby (Vice-Chair)

Ken Adler

Susan Anenberg

Chad Bailey

Larry Brockman

Linda Chappell

Benjamin DeAngelo

Rich Damberg

John Dawson

Neil Frank

Michael Geller

Gayle Hagler

Brooke Hemming

Lesley Jantarasami

Thomas Luben

John Mitchell

Jacob Moss

Contributing Authors

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

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

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

Jamie Bowers

Devin Hartman

Megan Melamed

Joseph Tikvart

Lourdes Morales

Joseph Dougherty

Additional Contributions to the Report

Project Support

Sonoma Technology, Inc.

Steve Brown

Chelsea Jennings

Marina Michaels

Jana Schwartz

Report Production

Stratus Consulting

Nimmi Damodaran

Joe Donahue

Editorial Support
Denise Mauzerall

Surabi Menon

Richard L. Poirot

Armistead (Ted) Russell

Michael Walsh

John Watson

Peer Review

C. Arden Pope, III (Chair)

Alberto Ayala

Michelle Bell

Kevin J. Boyle

Sylvia Brandt

Linda Bui

James J. Corbett

Ivan J. Fernandez

H. Christopher Frey

Jan Fuglestvedt

D. Alan Hansen

Joseph Helble

Mark Jacobson

Jonathan Levy

Council on Clean Air Compliance Analysis, Black Carbon Review Panel

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

EPA Science Advisory Board Staff

ii

Acknowledgments

Report to Congress on Black Carbon
Report to Congress on Black Carbon iii

Highlights

y Black carbon (BC) is the most strongly lightabsorbing component of particulate matter (PM), 

and is formed by the incomplete combustion of 

fossil fuels, biofuels, and biomass.

y 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. 

y BC contributes to the adverse impacts on human 

health, ecosystems, and visibility associated with 

PM2.5.

y BC influences climate by: 1) directly absorbing 

light, 2) reducing the reflectivity (“albedo”) 

of snow and ice through deposition, and 3) 

interacting with clouds. 

y 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. 

y Sensitive regions such as the Arctic and the 

Himalayas are particularly vulnerable to the 

warming and melting effects of BC. 

y 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. 

y 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. 

y 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. 

y 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. 

y Control technologies are available to reduce BC 

emissions from a number of source categories. 

y BC mitigation strategies, which lead to reductions 

in PM2.5, can provide substantial public health and 

environmental benefits. 

y 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. 

y 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. 

y 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.

y 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. 

y 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

Highlights

Report to Congress on Black Carbon
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

Report to Congress on Black Carbon v
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 PM2.5, 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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Report to Congress on Black Carbon vii
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

Table of Contents 

Report to Congress on Black Carbon ix
x Report to Congress on Black Carbon

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 m-2). . . . . . . . . . . . . . . . . . . 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
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. . . . . . . . . . . . . . . . . . 111

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 (μg m-3) 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 (μg/m3) in the United Kingdom, 1954-2005. . . 124

Figure 5-7b. Comparison of Ambient Black Smoke Measurements (μg/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

List of Figures

Report to Congress on Black Carbon xi
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 PM2.5 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 PM2.5 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

xii

List of Figures

Report to Congress on Black Carbon
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

List of Figures

Report to Congress on Black Carbon xiii
xiv Report to Congress on Black Carbon

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 PM2.5, 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 PM2.5 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
Report to Congress on Black Carbon x v

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

C1 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

COH Coefficient of Haze
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

EI 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 Grams per Break HorsepowerHour

GACC Global Alliance for Clean 

Cookstoves

GAW Aerosol Global Atmospheric Watch 

Aerosol Program

GCP Global Cost Potential

GDI Gasoline Direct Injection

GDP Global Damage Potential

GSFC Goddard Space Flight Center

Gg Gigagram (109 g = 1 kilotonne)

GHG Greenhouse Gas

GIZ German Agency for 

International Development 

(Deutsche Gesellschaft für 

Internationale Zusammenarbeit)

GLAS Geoscience Laser Altimeter 

System (NASA)

GTP Global Temperature Potential

GWP Global Warming Potential

H2O Water

HAP Hazardous Air Pollutant 

HC Hydrocarbon

HFC Hydrofluorocarbon 

HHK Hybrid Hoffman Kiln

HKHT Hindu-Kush Himalayan Tibetan 

Region

HULIS Humic-Like Substances

hPa Hectopascal (a unit of 

barometric pressure)

IAM Integrated Assessment Model

ICI Industrial, Commercial and 

Institutional (Boilers)

IEA International Energy Agency

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List of Acronyms

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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 OC:BC)

List of Acronyms

Report to Congress on Black Carbon xvii
OECD Organisation for Economic Cooperation and Development

OM Organic Matter

OMI Ozone Monitoring Instrument 

PAH Polycyclic Aromatic 

Hydrocarbon

PBL Planetary Boundary Layer

PCFV Partnership for Clean Fuels and 

Vehicles

PCIA Partnership for Clean Indoor Air

PDPF Partial Diesel Particulate Filters

PIC Product of Incomplete 

Combustion

PM Particulate Matter (for related 

terms, such as PM2.5 and PM10; 

see Glossary)

POM Polycyclic Organic Matter

RCP Representative Concentration 

Pathway

RHR Regional Haze Rule

RICE Reciprocating Internal 

Combustion Engine

RM Raman Microspectroscopy

RPO Regional Planning Organization

RWC Residential Wood Combustion

SAGE System for Assessing Aviation’s 

Global Emissions (FAA)

SCC Social Cost of Carbon

SCR Selective Catalytic Reduction

SEARCH SouthEastern Aerosol Research 

and Characterization

SEM Scanning Electron Microscopy

SFP Specific Forcing Pulse

SLCF Short-Lived Climate Forcer

SNV Netherlands Development 

Organization

SO2 Sulfur Dioxide

SOA Secondary Organic Aerosols

SPARC Stratospheric Processes and 

their Role in Climate (World 

Climate Research Programme)

SRES Special Report on Emissions 

Scenarios (IPCC)

STN Speciation Trends Network

STRE Surface Temperature Response 

per unit continuous Emission

TEM Transmission Electron 

Microscopy

TERP Texas Emissions Reduction Plan

TF HTAP Task Force on Hemispheric 

Transport of Air Pollution

Tg Teragram (1012 g = 1 

megatonne)

TOA Top of the Atmosphere

TOMS Total Ozone Mapping 

Spectrometer

TOR Thermal/Optical Reflectance

TOT Thermal/Optical Transmittance

μg/m3 Micrograms per cubic meter

ULSD Ultra-Low-Sulfur Diesel

UNEP United Nations Environment 

Programme

UNFCCC United Nations Framework 

Convention on Climate Change

USAID United States Agency for 

International Development

USDA United States Department of 

Agriculture

USGCRP United States Global Climate 

Change Research Program

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UV Ultraviolet

VA Vanillic Acid

VMT Vehicle Miles Traveled

VOC Volatile Organic Compound

VSBK Vertical Shaft Brick Kiln

VSL Value of a Statistical Life

W m-2 Watts per square meter (also 

W/m2)

WAIS West Antarctic Ice Sheet

WHO World Health Organization

WMO World Meteorological 

Organization

WRAP Western Regional Air 

Partnership

WTP Willingness to Pay

List of Acronyms

Report to Congress on Black Carbon xix
xx Report to Congress on Black Carbon

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). 
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). 

Glossary

Report to Congress on Black Carbon xxi
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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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. Filterbased 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.

Glossary

Report to Congress on Black Carbon xxiii
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

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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. 

Glossary

Report to Congress on Black Carbon xxv
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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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: PM10 and PM2.5.

PM10

Particulate matter with an aerodynamic diameter less than or equal to a nominal 10 micrometers. 

PM10 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[1,2,3cd]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).

Glossary

Report to Congress on Black Carbon xxvii
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 singlescattering 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.

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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.

Glossary

Report to Congress on Black Carbon xxix
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 precontrol (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 particulate 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

Glossary

Report to Congress on Black Carbon
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 H.R. 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 

y an inventory of the major sources of black carbon, 

y an assessment of the impacts of black carbon on 

global and regional climate, 

y 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, 

y an identification of the most cost-effective 

approaches to reduce black carbon emissions, 

and 

y 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 lightabsorbing 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 1
2

Executive Summary

Report to Congress on Black Carbon

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 

PM2.5; 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. 

other types of particles, including sulfates, nitrates 

and organic carbon (OC), generally reflect light. BC 

is a major component of “soot,” a complex lightabsorbing 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 (PM2.5). 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. 

Many countries have significantly higher 

PM2.5 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 PM2.5.

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 

Figure A. BC Emissions by Major Source Category. (Source: Lamarque et al., 2010 and U.S. EPA)
Executive Summary

Report to Congress on Black Carbon 3

4. BC influences climate through multiple 

mechanisms:

y Direct effect: BC absorbs both incoming and 

outgoing radiation of all wavelengths, which 

contributes to warming of the atmosphere and 

dimming at the surface. 

y Snow/ice albedo effect: BC deposited on snow and 

ice darkens the surface and decreases reflectivity, 

thereby increasing absorption and accelerating 

melting. 

y 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 

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)
4

Executive Summary

Report to Congress on Black Carbon

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 m-2). A recent UNEP/WMO 

assessment presented a narrower central range of 

+0.3 to +0.6 W m-2. These estimates are generally 

higher than the 2007 Intergovernmental Panel on 

Climate Change (IPCC) estimate of +0.34 (±0.25) W 

m-2. 

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 m-2 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 m-2 (with a 5 to 95% confidence 

range of -0.3 W m-2 to -1.80 W m-2). 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 m-2 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 m-2, with a best 

estimate of +0.6 W m-2; 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.
Executive Summary

Report to Congress on Black Carbon 5

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 copollutants; 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 longterm. 

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.

Figure C. BC Emissions, 2000, Gg. (Courtesy of Tami Bond, produced based on data from Bond et al., 2007)
6

Executive Summary

Report to Congress on Black Carbon

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 

PM2.5 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 PM2.5 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. 
Executive Summary

Report to Congress on Black Carbon 7

As illustrated in Figure D, the ideal emissions 

reduction strategies will also depend on a range of 

constraining factors, including: 

y Timing

y Location

y Atmospheric Transport

y Co-emitted Pollutants

y Cost

y Existing Regulatory Programs

y Implementation Barriers

y 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 PM2.5

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 

Figure D. Policy Framework for Black Carbon Mitigation Decisions. (Source: U.S. EPA.)
8

Executive Summary

Report to Congress on Black Carbon

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.

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. 

y 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. 

y 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 
Executive Summary

Report to Congress on Black Carbon 9

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. 

y Sensitive Regions: To address impacts in the 

Arctic, other assessments have identified the 

transportation sector (land-based diesel engines 

and Arctic shipping); residential heating (woodfired 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.

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 (PM2.5) 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.

Chapter 1

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 H.R. 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 

y an inventory of the major sources of black 

carbon, 

y an assessment of the impacts of black carbon on 

global and regional climate, 

y 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, 

y an identification of the most cost-effective 

approaches to reduce black carbon emissions, 

and 

y 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 

Report to Congress on Black Carbon 11
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 longlived 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 coemitted 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 PM2.5-

related human health impacts and other, nonclimate 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?

10.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.

12

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Report to Congress on Black Carbon
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 nearterm 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: 

y 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.

y There is virtual certainty that reducing primary 

PM will benefit public health.

y The Arctic, as well as alpine regions, may benefit 

more than other regions from reducing emissions 

of BC.

y 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.

y 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.

unece.org/env/lrtap/multi_h1.html.

Introduction

Report to Congress on Black Carbon 13
Tromsø 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:

y Addressing short-lived climate forcers such as 

black carbon, methane and ozone offers unique 

opportunities to slow Arctic warming in the near 

term.

y 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.

y 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.

y 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, and a number of 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. 

14

Chapter 1﻿

Report to Congress on Black Carbon
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 PM2.5 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 5 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 PM2.5 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 lightabsorbing 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.

Introduction

Report to Congress on Black Carbon 15
Appendix 1 provides further details regarding 

alternative definitions of BC and other lightabsorbing 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.

16

Chapter 1﻿

Report to Congress on Black Carbon
Black Carbon and Its Effects on 

Climate

2.1 Summary of Key Messages

y Black carbon (BC) is the most strongly lightabsorbing 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 lightabsorbing mixture that also contains organic 

carbon (OC). 

– Other carbon-based PM may also be lightabsorbing, 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.

y BC is always emitted with other particles and 

gases, such as sulfur dioxide (SO2), nitrogen 

oxides (NOx), 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. 

y Atmospheric processes that occur after BC is 

emitted, such as mixing, aging, and coating, can 

also affect the net influence of BC on climate.

y 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 

Chapter 2

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.

y 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. 

y 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). 

y 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 17
– 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. 

y 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. 

y 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. 

y BC also contributes to the formation of 

Atmospheric Brown Clouds (ABCs) and resultant 

changes in the pattern and intensity of 

precipitation. 

y 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. 

y There is currently no single metric that is widely 

accepted by the science and research community 

for this purpose.

y 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). 

y 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 nearterm 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. 

18

Chapter 2﻿

Report to Congress on Black Carbon
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 

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.

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 19
regional climate, highlighting the effect of BC in 

sensitive regions such as the Arctic and other snowand 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; Moosmüller 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 lightabsorbing 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 lightabsorption 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 (μm), which aggregate 

to form particles of larger sizes (0.1 to 1 μm) (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.

1 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).

Figure 2-2. BC Images. (a) High resolution transmission electron microscopy (TEM) image of a BC spherule 

(Pósfai and Buseck, 2010). (b) TEM image of a representative soot particle. Freshly emitted soot particles are 

aggregates of soot spherules (Alexander et al., 2008).

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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 OC:BC 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 nonorganic, 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 OC:BC 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; Gelencsér, 2004; Andreae and 

Gelencsér, 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.

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 

3During 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).

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. 

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)

(a) (b) (c)

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 21
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 incoming 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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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 

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 CO2. 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:

y Gas phase fuels (e.g., natural gas) can be readily mixed with oxygen, which reduces the emission of carbonaceous 

particles. 

y 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. 

y 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; Pósfai 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, cooler form of 

combustion which occurs as oxygen directly attacks the surface of heated solid fuel. The smoke that appears is lightcolored, 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.

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 23
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.

balls” or “carbon spheres,” range in diameter from 

0.03 and 0.5 μm (Pósfai 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 

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)

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)

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There is also a significant BC component in (or rather 

on) coarse particles (PM10-2.5 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 Gelencsér, 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 

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 25
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 (HFCs) 

have lifetimes that range from as short as a year 

for some of the HFCs to as long as 50,000 years 

for tetrafluoromethane (CF4), a perfluorocarbon 

(Forster et al., 2007).6 The 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 

6 Several 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.

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 Days to weeks Up to millenniae

Distribution of atmospheric 

concentrations

Highly variable both geographically and 

temporally, correlating with emission sources Generally uniform across globe

Direct radiative properties Absorbs all wavelengths of solar radiation Absorbs only thermal infrared radiation

Global mean radiative forcing

+0.34 to 1.0 W m-2 direct forcinga

+0.05 W m-2 (snow/ice albedo forcing)b

± ? (cloud interactions)c

Net effect: uncertain, but likely warming

+1.66 (±0.17) W m-2 (f)

Cloud interactions

Multiple cloud interactions that can lead to 

warming or cooling (typically cooling), as well as 

effects on precipitation

Increases cloud droplet acidity

Surface albedo effects

Contributes to accelerated melting of snow/ice 

and reduces reflectivity by darkening snow and 

ice, enhancing climate warming

No direct surface albedo effects

Contribution to current global 

warming

Likely third largest contributor (after CO2 and 

CH4), but large uncertaintyd Largest contributor

Dimming Contributes to surface dimming No direct effects on surface dimming

Acidification/fertilization No ocean acidification/fertilization effects 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.

f Forster et al. (2007).

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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 BanWeiss 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 

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 27
decades from now,7

 which is mainly driven by past, 

current, and future emissions of persistent species 

like CO2 combined with emissions of BC at that future 

date. Therefore, only sustained reductions in longlived GHGs can avert long-term climate change. 

In order to avoid (rather than delay) reaching 

temperature thresholds, long-lived GHG emissions 

reductions are necessary, but future BC reductions 

in conjunction with these GHG reductions can “shave 

the peak” off the temperature change and help 

avoid crossing these thresholds (possibly avoiding 

key non-linear climate impacts, often referred to 

as “tipping points”) (Lenton et al., 2009). Figure 2-8 

demonstrates this visually: reductions in emissions 

of CH4 and BC [the specific measures involved are 

described in more detail in the UNEP report (UNEP 

and WMO, 2011a)] lead to a near-term reduction 

in the rate of warming, but in the long run serve 

only to delay any given temperature change. While 

CO2 reduction measures do little to slow nearterm warming (in this case, in part because they 

are accompanied by simultaneous reductions of 

7As shown in Figure 2-7, the temperature response to an immediate 

reduction in BC can last a couple of decades – due mainly to inertia 

in the climate because of how the ocean takes up heat. However, 

not included in this figure is the possibility that these temperature 

changes can have feedbacks which might persist for even longer 

– for example, higher temperatures reduce the rate at which CO2

is mixed into the ocean, so a change in BC emissions today could 

change CO2 concentrations, and therefore temperatures, for many 

decades.

cooling aerosols), they are necessary for long-term 

stabilization of the climate. Reducing short-lived 

forcers in conjunction with CO2 enables both a 

reduction in near-term warming (due to near‐term 

BC and CH4 emissions reductions) and a decrease in 

peak warming (due to future BC and CH4 reductions 

combined with CO2 reductions over the entire time 

period).

These differences between BC and GHGs have 

significant implications for BC mitigation decisions. 

Specifically, the effectiveness of a given mitigation 

effort depends on the timing and location of the 

emissions; the atmospheric processes, transport, and 

deposition rates of the emissions from the specific 

sources; the underlying surface (e.g., ice and snow); 

and the presence of co-pollutant emissions. This 

constitutes a significant difference from long‐lived 

GHGs where the precise timing and location of 

emissions (or emissions reductions) does not matter 

significantly with respect to the climate impact. 

2.5 The Role of Co-Emitted Pollutants 

and Atmospheric Processing

As discussed in section 2.3, BC is never emitted 

into the atmosphere in isolation. Rather, it is always 

emitted as part of a mixture of particles and gases 

during the combustion process. The composition 

of this mixture can vary significantly, depending on 

combustion conditions and fuel type. BC is generally 

Figure 2-7. Climate Response 

to Emissions of Pollutants with 

Different Lifetimes. Figure 

illustrates forcing and temperature 

response to pulse emissions of 

three hypothetical substances with 

different atmospheric lifetimes 

(Column 1 illustrates short-lived 

species like BC, Column 2 illustrates 

species with medium atmospheric 

lifetime like methane, and Column 

3 illustrates a long-lived species 

like CO2). In the middle row, which 

illustrates global average forcing, 

the dashed red trace shows forcing 

value times 50 to show more detail 

in the time dependence. (Bond et 

al., 2011)

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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 OC:BC 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; 

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.) (UNEP and WMO, 2011a)

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 29
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 

OC:BC 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 

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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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 al., 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. 

a Fresh BC is produced primarily during flaming combustion. 

b 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. 

c Emitted directly as a byproduct of combustion, or formed through the oxidation of SOx

 or NOx. 

d In the exhaust gases of solid fuel fires, low volatility BrC and other organic compounds can condense on BC particles. In 

the ambient atmosphere, low volatility organics produced by oxidation of VOCs can also condense on BC. 

e Forms when high particle concentrations lead to the coagulation of multiple particles. 

f Forms by condensation of water vapor onto acidic organic (carbon-based) and inorganic particles. 

g Forms when complex particles undergo the humidification and drying cycles characteristic of cloud formation and 

evaporation. 

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 31
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 m-2). 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 “topof-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. 

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 longlived 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. 

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 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.

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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-2 for 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.), 

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 33
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 

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 CH4 from 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 CH4 emissions. Uncertainty in 

direct forcing is ±0.25 W m-2 for BC and ±0.20 W m-2 for both OC and SO2. The range of forcing for the cloud 

albedo effect is -1.8 to -0.3 W m-2. (Adapted from Figure 2.21 of Forster et al., 2007) 

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Report to Congress on Black Carbon
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 m-2, with a 

central value of 0.6 W m-2 (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,9 and the current lack of quantitative 

9 The IPCC estimated the net forcing due to all aerosols as 

-1.2 W m-2 (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 m-2), while Hansen et al. (2011) argue the 

effect is more strongly negative (-1.6 W m-2 ±0.3)

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) 

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 35
estimates for all processes. In light of these 

constraints, they argued that net effective BC forcing 

is unlikely to exceed 1 W m-2.

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 m-2. 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 

m-2. 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 m-2. Additional work 

Forcing Range (W/m2) Central Values (W/m2) Mean Value (W/m2)

Direct 0.2 to 0.9 0.3 to 0.6 0.45

Semi-direct and indirect -0.6 to 0.5 -0.4 to 0.4 0.0

Deposition

(0.01 to 0.08)

corrected for efficacy (x5)

0.05 to 0.4

(0.01 to 0.05)

0.05 to 0.25

0.03

0.15

Total BC

-0.3 to 1.8 0.0 to 1.3

0.0 to 1.0 with constraints on total 

aerosol forcing

0.5

0.6 including efficacy of deposition

Table 2-3. Summary of UNEP/WMO Assessment Estimates of Radiative Forcing Effects of BC. (Adapted 

from UNEP and WMO, 2011a, Table 3-1)

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Report to Congress on Black Carbon
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 m-2. 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 m-2, 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 m-2.

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 m-2 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 m-2, 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.

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, postindustrial, present day). Note: 

BB = open biomass burning; 

BF = biofuels; and FF = fossil 

fuels. (Source: U.S. EPA)

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 37
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 m-2 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 m-2 (Chung and Seinfeld, 2002). For the Southern 

Hemisphere, Chung and Seinfeld (2005) estimate a 

range of +0.15 to +0.23 W m-2. 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 m-2. 

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 m-2) of BC from All Sources, simulated with the Community 

Atmosphere Model. (Bond et al., 2011)

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Report to Congress on Black Carbon
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 “mixedphase 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 m-2.

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 

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 Cloud Types 

Affected Process Net Effect Potential Magnitude

Scientific 

Understanding

Cloud albedo effect 

(first indirect effect) All clouds Smaller cloud particles reflect more solar radiation Cooling Medium Low

Cloud lifetime effect 

(second indirect 

effect)

All clouds Smaller cloud particles decrease 

precipitation so increase lifetime Cooling Medium Very low

Semi-direct effect All clouds

Absorption of solar radiation 

by absorbing aerosols changes 

atmospheric stability and cloud 

formation

Cooling or 

warming Small Very low

Glaciation indirect 

effect

Mixed-phase 

clouds

An increase in ice nuclei due 

to some aerosols increases 

precipitation

Warming Medium Very low

Thermodynamic 

effect

Mixed-phase 

clouds

Smaller cloud droplets delay 

freezing causing super-cooled 

clouds to extend to colder 

temperatures

Cooling or 

warming Medium Very low

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 39
as mineral dust) serving as ice nuclei in a supercooled 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 

m-2, 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 m-2.

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 mixedphase 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 m-2), but with a 

central range of -0.4 to +0.4 W m-2. 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.

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Report to Congress on Black Carbon
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 m-2 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 m-2, with a probable range 

of 0 to +0.1 W m-2 (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 m-2, of 

which +0.033 W m-2 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 m-2. Bond et 

al. (2011) estimated a global forcing of +0.047 W m-2, 

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 m-2. In 

line with these more recent studies, the UNEP/WMO 

assessment (UNEP and WMO, 2011a) estimated 0.05 

W m-2, with a range of 0.01 to 0.10 W m-2, 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) 

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)

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 41
would be consistent with a total albedo feedback on 

the order of +0.45 W m-2. 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 fossilfuel 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 m-2 in the Tibetan plateau, 

with instantaneous forcings13 of up to +20 W m-2 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 

m-2 in the Arctic compared to +0.3 W m-2 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.

Figure 2-15. Snow and Ice Albedo Forcing by BC, simulated with the Community Atmosphere Model. 

(Bond et al., 2011) 

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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 BC and OC 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)

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)

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 43
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 m-2 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 m-2 (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 m-2 

(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 

m-2 (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 

Figure 2-18. Direct Forcing by OC from All Sources, simulated with the Community Atmosphere Model. 

(Bond et al., 2011)

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Report to Congress on Black Carbon
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 m-2, in contrast to a 

calculated BC direct forcing of 

+0.9 W m-2. 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

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 

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 

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. (Unger et al., 2010)

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 45
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 m-2 (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.

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Report to Congress on Black Carbon
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.

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 47
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—

Figure 2-20. Spatial Distribution of Change in Mean Snow Water Equivalent (SWE, mm) for March. (Qian et al., 2009)

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Report to Congress on Black Carbon
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 

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)

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 49
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 m-2, with about 

-3.4 W m-2 from the direct effect of aerosols (roughly 

half of which is attributed to BC) and the remaining 

-1 W m-2 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.) 

Effects Sign of Change in 

Surface Dimming

Potential 

Magnitude

Sign of Change in 

Precipitation

Potential 

Magnitude

Cloud albedo effect Positive Medium NA NA

Cloud lifetime effect Positive Medium Negative Small

Semi-direct effect Positive Large Negative Large

Glaciation indirect effect Negative Medium Positive Medium

Thermodynamic effect Positive or negative Medium Positive or negative Medium

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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 

Figure 2-22. Surface Dimming by Anthropogenic Aerosols (W m-2). (Adapted from Chung et al., 2005, 

and Ramanathan and Carmichael, 2008)

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 51
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 0o and 20o

N) and decreased precipitation just south of the 

equator (between 0o and 20o 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 highlevel 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 

Niño/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 northsouth 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 m-2 in the 

spring, +0.66 W m-2 in the summer, +0.16 W m-2

in the fall, and only 0.09 W m-2 in the winter. Snow 

albedo forcing in the Arctic was calculated to add 

an additional +0.53 W m-2 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 

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Report to Congress on Black Carbon
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 

Flanner 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 

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).

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 53
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 

Flanner et al. (2007), land snowmelt rates north of 

50oN 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 

65oN 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 40oN 

and 50oN latitude, emissions of BC from between 

50oN and 60oN latitude had about three times as 

much forcing impact in the Arctic on a per-ton basis. 

In addition, emissions from north of 60oN 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 

Table 2-6. Arctic Temperature Impacts from Emissions of BC from Different Sectors. 

Scale of 

Physical Impact

Estimated Forcing or 

Temperature Change

Source of 

Emissions Aerosols Model Reference Notes

Arctic 0.5 – 1.4° C 100% of FF, 

BF, BB BC, OC GISS-ER Shindell and Faluvegi (2009)

Aerosol indirect 

included “crudely.”

Arctic 0.5 – 1.6° C 100% of FF, 

BF, BB BC, OC NCAR-CAM3 and SNICAR

Flanner et al. 

(2007)

Range results 

from using fire 

frequencies in high 

year globally (1998) 

and low year (2001).

Arctic 1.2° C 100% of FF BC, OC, minor 

inorganics

GATORGCMOM Jacobson (2010)

Arctic 1.7° C 100% of FF, BF

Same as 

above plus 

CO2, CH4

Same as 

above Jacobson (2010)

Arctic Indirect > 0 100% of All 

sources

All aerosols

Idealized 

calculations, 

observations

Mauritsen et al. 

(2010)

Indirect only, no 

direct or snow 

albedo effect.

FF = fossil fuel; BF = biofuel; BB = biomass burning. 

54

Chapter 2﻿

Report to Congress on Black Carbon
emissions from regions below 50oN, and most of the 

remainder was derived from emissions from regions 

between 50oN and 60oN.

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-HimalayanTibetan (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 m-2 with short-term forcing of up to 20 

W m-2 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 0.93oC 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). 

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 55
2.6.6 Summary of BC Impacts in Key 

Regions

As described in the previous sections, the climaterelated 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 U.S. Asia Arctic

Radiative Forcing 

Effects

•	 Estimates of direct radiative forcing 

of BC over the United States range 

from 0.1 to 0.7 W m-2. 

•	 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 m-2, with 

local instantaneous forcing up to 20 

W m-2.

•	 Springtime Arctic forcing has 

been estimated to be 1.2 W 

m-2 (direct) and 0.53 W m-2

(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.

56

Chapter 2﻿

Report to Congress on Black Carbon
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 

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.

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 57
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?

Metric Type Climate Impact Baseline 

Forcer

Emissions 

Type

Spatial 

Scale

Includes Rate 

of Change?

GWP (Global Warming Potential) Integrated radiative 

forcing CO2 Pulse Global No

GTP-pulse (Global Temperature Potential) Temperature CO2 Pulse Global No

GTP-sustained Temperature CO2 Sustained Global No

STRE (Surface Temperature Response per 

unit continuous Emission) Temperature CO2 Sustained Global No

SFP (Specific Forcing Pulse) Energy Joules/gram Pulse Global or 

regional No

Cost-effectiveness Metrics (e.g., Manne 

and Richels, 2001, Global Cost Potential) Mainly temperature CO2or $ value

Optimal 

emissions 

calculation

Global Optional

Value of Damages (e.g., Social Cost of 

Carbon, Global Damage Potential)

Range of climate 

damages $ value Pulse Global Limited

Table 2-8. Examples of Commonly Used Metrics for GHGs. 

58

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Report to Congress on Black Carbon
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 over time). 

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 welldefined metric is the GWP. The definition of the GWP 

by the IPCC (2007) is

“An index, based upon radiative properties of wellmixed 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 1).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-2 yr g-1. 

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 59
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/climatechange/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 costeffectiveness 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 2oC, 

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 2o 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 

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 61
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 

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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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 longlived 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 coemissions 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 

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.

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 63
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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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 longlived 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 wellmixed 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. 

Black Carbon and Its Effects on Climate

Report to Congress on Black Carbon 65
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 snowand 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-relevant 

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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Black Carbon Effects on Public 

Health and the Environment

3.1 Summary of Key Messages

y 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.

y 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. 

y 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 PM2.5; 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.

y 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 

Chapter 3

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. 

y PM2.5, 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 nonclimate 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
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 longterm [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 

PM2.5 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 PM2.5 constituents 

observed in different geographical regions as well as 

regional differences in PM2.5‐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 PM2.5 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 PM10-2.5 mass 

Exposure Outcome Causality Determination

Short-term exposure to PM2.5

Cardiovascular Effects Causal

Respiratory Effects Likely to be causal

Mortality Causal

Long-term exposure to PM2.5

Cardiovascular Effects Causal

Respiratory Effects Likely to be causal

Mortality Causal

Reproductive and Developmental Effects Suggestive

Cancer, Mutagenicity, and Genotoxicity Suggestive

Table 3-1. Summary of Causal Determinations for Exposure to PM2.5 from 2009 PM ISA. (Source: U.S. EPA)

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Chapter 3﻿

Report to Congress on Black Carbon
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 PM10-2.5 and health effects, 

the current evidence, primarily from epidemiological 

studies, indicates that short-term exposure to PM10-2.5

is associated with effects on both the cardiovascular 

and respiratory systems. However, variability in the 

chemical and biological composition of PM10-2.5, 

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 thermaloptical 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 DNA 

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 DNA 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; Rückerl 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 BNP) (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.

Black Carbon Effects on Public Health and the Environment

Report to Congress on Black Carbon 69
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)

Reference 

Number in 

Figure 3-1

Study Location Health Outcome Metric Representative Concentration (μg/m3

)

Exposure 

Assessmenta

Selected Effect Estimates

(95% CI)b

Short-term Exposure Studies

—

McCracken et al. (2010) 

Boston, MA Telomere Length BC Mean (annual avg): 0.32 Annual outdoor home 

concentration estimates from spatiotemporal model

Association with leukocyte telomere length:-7.6% (-12.8%, -2.1%)

1.

Delfino et al. (2010)

Los Angeles, CA

Blood Pressure BC Mean (24-h avg): 1.67 Hourly outdoor 

home air-pollutant concentrations

Change in BP: SBP: 0.22 (-0.65, 1.09)DBP: 0.36 (-0.11, 0.83)

2.

Mordukhovich et al. (2009) 

Boston, MA Blood Pressure BC Mean (7-day moving average): 

1.10

Continuous measurements from single monitor averaged by hour before BP measurement

Change in BP: SBP: 1.46 (0.10, 2.82)DBP: 0.87 (0.15, 1.59)

3.

Wilker et al. (2010)

Boston, MA Blood Pressure BC Mean (7-day moving average): 

0.98

Continuous measurements from single monitor averaged by hour before BP measurement

Change in BP: SBP: 3.52 (2.77, 4.26)DBP: 2.72 (2.31, 3.12)

4.

Adar et al. (2007)

St. Louis, MO HRV BC Median (5-min avg periods): 

0.285-2.911 (range across microenvironments)

Two portable carts containing continuous sampling instrumentation

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) 

5.

Chuang et al. (2008)

Boston, MA HRV BC Median (24-h avg): 0.79 Continuous 

measurements from single monitor

Relative risk for ST-segment Depression ≤1 mm:1.50 (1.19-1.89)

6.

Gold et al. (2005)

Boston, MA HRV BC Median (12-h avg): 1.14 Continuous 

measurements from single monitor

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)

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Reference 

Number in 

Figure 3-1

Study Location Health Outcome Metric Representative Concentration (μg/m3

)

Exposure 

Assessmenta

Selected Effect Estimates

(95% CI)b

7.

Park et al. (2005)

Boston, MA HRV BC Mean (24-h avg): 0.92 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)

8.

Schwartz et al. (2005)

Boston, MA HRV BC Mean (24-h avg): 1.2 Continuous 

measurements from single monitor

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)

9.

Delfino et al. (2011)

Los Angeles, CA

HRV BC Mean (24-h avg): 1.67 Hourly outdoor 

home air-pollutant concentrations

Odds ratio for ST-segment Depression ≤1 mm:2.07 (1.30, 3.29)

10.

Park et al. (2008)

Boston, MA Plasma Total 

Homocysteine

BC Mean (24-h avg): 0.99 Continuous 

measurements from single monitor

Percent change in total homocysteine:3.13 (0.76, 5.55)

11.

Ren et al. (2010)

Boston, MA Plasma Total 

Homocysteine

BC Mean (7-day moving average): 

0.99

Continuous measurements from single monitor

Percent change in total homocysteine:0.68 (-0.46, 1.81)

—

Baccarelli et al. (2009) 

Boston, MA Blood DNA 

Methylation

BC Mean (24-h avg): 0.89 Continuous 

measurements from single monitor

Coefficient for effect on methylation:LINE-1: -0.09 (-0.15, -0.02)Alu: -0.02 (-0.08, 0.05)

12.

Baja et al. (2010) 

Boston, MA Ventricular 

Repolarization

BC Mean (1-h avg): 1.08 Continuous 

measurements from single monitor

Change in mean QTc (msec):1.89 (-0.16, 3.93)

13.

Dockery et al. (2005)

Boston, MA Arrhythmia BC Median (2-day avg): 0.98 Continuous 

measurements from single monitor

Association with ventricular arrhythmias:Recent arrhythmia > 3 days: 1.02 (0.83, 1.24)Recent arrhythmia <3 days: 1.74 (1.28, 2.37)

14.

Rich et al. (2005) 

Boston, MA Arrhythmia BC Median (24-h avg): 0.94 Continuous 

measurements from single monitor

Odds ratio for ventricular arrhythmias:0.93 (0.74, 1.18)

15.

Rich et al. (2006) 

Boston, MA Arrhythmia BC Median (24-h avg): 0.94 Continuous 

measurements from single monitor

Odds ratio for paroxysmal atrial fibrillation:1.46 (0.67, 3.17)

Black Carbon Effects on Public Health and the Environment

Report to Congress on Black Carbon 71
Reference 

Number in 

Figure 3-1

Study Location Health Outcome Metric Representative Concentration (μg/m3

)

Exposure 

Assessmenta

Selected Effect Estimates

(95% CI)b

16.

Zanobetti et al. (2009)

Boston, MA Arrhythmia BC Median (6-h avg): 

Ambient: 0.72Indoor: 0.41Outdoor: 0.50

Continuous ambient measurements from single monitor; indoor and outdoor measured continuously at participants’ homes

Odds ratio for maximum T-wave alternans ≤ 26 μV:1.42 (1.19, 1.69)

17.

Rich et al. (2004)

Vancouver, Canada

Arrhythmia EC Mean (24-h avg): 0.8 Continuous 

measurements from single monitor

Odds ratio for defibrillator discharge:1.06 (0.87, 1.33)**estimated from graph

18.

Sorensen et al. (2003)

Copenhagen, Denmark

Blood Markers of Coagulation and Inflammation

Carbon 

Black

Median (24-h avg): 8.1 (10-6/m)

Personal exposure Association with plasma proteins:

4.1% increase in plasma proteins per 1 x 10-5/m increase in personal CB exposure

19.

Dubowsky et al. (2006) 

St. Louis, MO Blood Markers of 

Coagulation and Inflammation

BC Mean (24-h avg): 0.9 Continuous 

measurements from single monitor

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)

20.

Ruckerl et al. (2006)

Erfurt, Germany

Blood Markers of Coagulation and Inflammation

EC Mean (24-h avg): 2..6 Continuous 

measurements from single monitor

Odds ratio for increase in blood marker above 90th percentile:CRP: 1.3 (0.7, 2.4)ICAM-1: 2.6 (1.7, 3.8)

21.

Delfino et al. (2008)

Los Angeles, CA

Blood Markers of Coagulation and Inflammation

BC Mean (24-h avg): 2.00 Outdoor home 

measurements

Coefficient for association:CRP (ng/ml): 585.61 IL-6 (pg/ml): 0.48sTNF-RII (pg/ml): 135.15sP-selectin (ng/ml): 1.99Cu, Zn-SOD (U/g Hb): -187.95

22.

Delfino et al. (2009)

Los Angeles, CA

Blood Markers of Coagulation and Inflammation

BC Mean (24-h avg): 1.59 – 1.76 Hourly outdoor home air 

pollutants 

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-α (pg/ml): 0.02 (-0.06, 0.10)GPx-1 (U/g Hb): -0.47 (-0.97, 0.03)

23.

O’Neill et al. (2007)

Boston, MA` Blood Markers of 

Coagulation and Inflammation

BC Mean (24-h avg): 1.1 Continuous 

measurements from single monitor

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)

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Reference 

Number in 

Figure 3-1

Study Location Health Outcome Metric Representative Concentration (μg/m3

)

Exposure 

Assessmenta

Selected Effect Estimates

(95% CI)b

24.

Wellenius et al. (2007) 

Boston, MA Blood Markers of 

Coagulation and Inflammation

BC Mean (24-h avg): 0.73 Continuous 

measurements from single monitor

“No significant associations observed between [BC] and BNP levels at any of the lags examined.”

25.

Zeka et al. (2006) 

Boston, MA Blood Markers of 

Coagulation and Inflammation

BC Mean (2-day avg): 0.77 Continuous 

measurements from single monitor

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)

26.

Peters et al. (2001) 

Boston, MA ED visits 

and hospital admissions for CVD

BC Mean (24-h avg): 1.35 Continuous 

measurements from single monitor

Odds ratio for MI:1.21 (0.87, 1.70)

27.

Zanobetti and Schwartz (2006) 

Boston, MA ED visits 

and hospital admissions for CVD

BC Median (24-h avg): 1.15 Continuous 

measurements from single monitor

Percent change in MI admissions:8.34 (0.21, 15.82)

28.

Bell et al. (2009) 

106 U.S. Counties

ED visits and hospital admissions for CVD

EC Mean (24-h avg): 0.72 County-wide averages Percent increase in health effect 

estimate:25.8 (4.4, 47.2)

29.

Peng et al. (2009) 

119 U.S. Counties

ED visits and hospital admissions for CVD

EC Median (24-h avg): 0.58 County-wide averages Percent increase in CVD hospital 

admissions:0.72 (0.43, 1.01)

30.

Sarnat et al. (2008)

Atlanta, GA ED visits 

and hospital admissions for CVD

EC Mean (24-h avg): 1.4-1.7 Positive matrix 

factorization applied to measurements from single monitor

Relative risk of ED visit for CVD:1.025 (1.014, 1.036)

31.

Tolbert et al. (2007) 

Atlanta, GA ED visits 

and hospital admissions for CVD

EC Mean (24-h avg): 1.6 Continuous 

measurements from single monitor

Relative risk of ED visit for CVD:1.015 (1.005, 1.025)

32.

Ito et al. (2011) New York, NY ED visits 

and hospital admissions for CVD

EC Mean (24-h avg): 1.13 Average of continuous 

measurements from three monitors

Percent excess risk:1.4 (0.1, 2.7)

Black Carbon Effects on Public Health and the Environment

Report to Congress on Black Carbon 73
Reference 

Number in 

Figure 3-1

Study Location Health Outcome Metric Representative Concentration (μg/m3

)

Exposure 

Assessmenta

Selected Effect Estimates

(95% CI)b

Long-term exposure studies

33.

O’Neill et al. (2005)

Boston, MA Endothelial 

dysfunction

BC Mean (24-h avg): 1.0 Continuous 

measurements from single monitor

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)

34.

Madrigano et al. (2010) 

Boston, MA Endothelial 

dysfunction

BC Mean (24-h avg): 0.84 Continuous 

measurements from single monitor

Percent change in blood markers:sVCAM-1: 4.52 (1.09, 7.96)sICAM-1: -1.37 (-3.89, 1.15)

35.

Alexeeff et al. (2011) 

Boston, MA Markers of 

inflammation and endothelial response

BC Mean (24-h avg): 0.42 Validated spatiotemporal (land use 

regression) model

Percent change in blood markers:4 wk exposure, sVCAM-1: 1.00 (-0.65, 2.67)4 wk exposure, sICAM-1: 1.50 (0.22, 2.80)8 wk exposure, sVCAM-1: 1.20 (-0.58. 3.02)8 wk exposure, sICAM-1: 1.58 (0.18, 3.00)12 wk exposure, sVCAM-1: 1.26 (-0.58, 3.14)12wk exposure, sICAM-1: 1.49 (0.04, 2.95)

36.

Gan et al. (2011) Vancouver, BC CHD 

hospitalizations

BC Mean (24-h avg): 1.19 High resolution land use 

regression model

Relative Risk of CHD hospitalization:1.01 (1.00, 1.03)

a A more complete description or evaluation of the BC monitoring method is beyond the scope of this summary, and may include surrogate measurements.b 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-1=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-α=tumor necrosis factor-alpha; GPx-1=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.

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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 PM10

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, 

Black Carbon Effects on Public Health and the Environment

Report to Congress on Black Carbon 75
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.

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) 

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Table 3-3. Summary of Epidemiological Studies of BC and Respiratory Health Outcomes. (Source: U.S. EPA)

Study Location Health Outcome Metric

Representative 

Concentration

 (μg/m3

)

Exposure Assessmenta Selected Effect Estimates (95% CI)

Clark et al. (2010) British Columbia Development of 

childhood asthma

BC Mean (24-h avg):

1.34

Land Use Regression (LUR) Model

Asthma risk due to average exposuresIn utero exposure: 1.08 (1.02, 1.15)First-year exposure: 1.14 (1.01, 1.29)

Delfino et al. (2006) Los Angeles, CA Exhaled nitric oxide 

(biomarker of airway inflammation)

EC Mean (24-h avg):

0.71-1.61

Personal exposure and continuous measurements from central site monitors

Association between EC and exhaled NO:Personal exposure: 0.72 (0.32, 1.12)Central site exposure: 1.38 (0.15, 2.61)

Jansen et al. (2005) Seattle, WA Exhaled nitric oxide 

(biomarker of airway inflammation)

BC Mean (24-h avg):

Central site: 2.01Indoor: 1.34Personal: 1.64

Personal exposures, indoor monitors, and central outdoor monitoring site

Association between BC and 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)

Kim et al. (2004) San Francisco, CA Bronchitis symptoms 

and asthma

BC Mean (24-h avg):

0.8

BC measured at 10 school sites

Odds ratios of respiratory illness by schoolbased BC concentration:Bronchitis: 1.04 (1.00, 1.08)Asthma: 1.07 (0.98, 1.17)

Mann et al. (2010) Fresno, CA Wheeze among 

asthmatic children

EC 

(estimated 

from BC)

Median (24-h avg): 1.3

Continuous measurements from single monitor

Odds ratio of EC and wheeze:1.12 (0.97, 1.30)

McConnell et al. (2003)

12 Southern California communities

Bronchitic symptoms among asthmatic children

EC Mean (24-h avg): 

0.71

Annual averages computed from 2-week averages measured in each community

Bronchitic symptoms as a function of the 4-yr avg EC concentration:1.64 (1.06, 2.54)

Patel et al. (2010) New York City, NY Respiratory 

symptoms among asthmatics and nonasthmatics

BC Median (24-h avg): 

0.49-2.4

BC measured at 4 high schools

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)

Spira-Cohen et al. (2011)

South Bronx, NY Lung function and 

respiratory symptoms among asthmatic children

EC Mean (24-h avg): 

1.9

Personal and outdoor school site monitoring at 4 elementary schools

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)

Black Carbon Effects on Public Health and the Environment

Report to Congress on Black Carbon 77
Study Location Health Outcome Metric

Representative 

Concentration

 (μg/m3

)

Exposure Assessmenta Selected Effect Estimates (95% CI)

Suglia et al. (2008) Boston, MA Lung function among 

women

BC Mean (predicted 

annual): 0.62

Local BC levels estimated by spatiotemporal land-use regression model

Effect estimates for change in lung function:FEV1: -1.09 (-2.5, 0.3)FVC: -0.62 (-1.9, 0.6)FEF25-75%: -3.03 (-5.8, -0.3)

Bell et al. (2009) 106 U.S. Counties ED visits and hospital 

admissions for respiratory effects

EC Mean (24-h avg): 

0.72

County-wide averages Percent increase in health effect estimate:

511 (80.7, 941)

Ostro et al. (2009) 6 CA Counties ED visits and hospital 

admissions for respiratory effects

EC Mean (24-h avg): 

1.0

County-wide averages Excess risk per IQR:

Ages <19: 5.4 (0.8, 10.3)Ages <19, Cool season only: 6.8 (-0.2, 14.2)

Peng et al. (2009) 119 U.S. Counties ED visits and hospital 

admissions for respiratory effects

EC Median (24-h avg): 

0.58

County-wide averages Percent increase in respiratory hospital 

admissions:0.43 (-0.02, 0.85)

Mohr et al. (2008) St. Louis, MO ED visits and hospital 

admissions for respiratory effects

EC Median (24-h avg): 

~0.6-0.7

Continuous measurements from single monitor

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)

Sarnat et al. (2008) Atlanta, GA ED visits and hospital 

admissions for respiratory effects

EC Mean (24-h avg): 

1.4-1.7

Positive matrix factorization applied to measurements from single monitor

Relative risk of ED visit for Resp:0.996 (0.988, 1.003)

Tolbert et al. (2007) Atlanta, GA ED visits and hospital 

admissions for respiratory effects

EC Mean (24-h avg): 

1.6

Continuous measurements from single monitor

Relative risk of ED visit for respiratory effects:0.996 (0.989, 1.004)

Zanobetti and Schwartz (2006)

Boston, MA ED visits and hospital 

admissions for respiratory effects

BC Median (24-h avg): 

1.15

Continuous measurements from single monitor

Percent change in pneumonia ED visits:11.71 (4.79, 17.36)

a A more complete description or evaluation of the BC monitoring method is beyond the scope of this summary, and may include surrogate measurements.b NO=nitric oxide; PEF=peak expiratory flow; FEV =forced expiratory volume in 1 second; FVC=forced vital capacity; FEF 1 25-75%=difference between the 25th and 75th percentile of forced 

expiratory flow; ED=emergency department.

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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 

Number in 

Figure 3-1

Study Location Health Outcome Metric Representative Concentration (μg/m3

)

Exposure 

Assessment1 Selected Effect Estimates (95% CI)

Short-term Exposure Studies

— Cakmak et al. (2011) 3 urban areas in 

Chile

All-cause mortality

EC Mean (24-h avg): 

2.69 – 5.37

Continuous measurements from 3 monitors (one in each urban area)

Mortality risk ratio:All-cause: 1.084 (1.067, 1.100)≤64 yrs old: 1.052 (1.019, 1.085)≥85 yrs old: 1.137 (1.103, 1.173)

37. Ito et al. (2011) New York City, 

NY

CVD mortality

EC Mean (24-h avg): 1.13 Continuous 

measurements from 3 monitors 

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)

— Lippmann et al. 

(2006)

90 U.S. cities and Hong Kong

All-cause mortality

EC (Not reported) Speciation data from 

U.S. EPA AQS from 2000-2003

“Elevated but nonsignificant increases >0.21 [in health effect estimate] were associated with EC”

38. Ostro et al. (2007) 6 California 

counties

All-cause, CVD and respiratory mortality

EC Mean (24-h avg): 

0.966

Continuous measurements across monitors

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)

39. Zhou et al. (2011) Seattle, WA and 

Detroit, MI

All-cause, CVD and respiratory mortality

BC (as a 

surrogate 

index 

of EC in 

the filter 

samples)

Median (24-h avg): 0.52-0.71

Continuous measurements from 1 monitor in each city

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

— Lipfert et al. (2006) 187 U.S. 

Counties

All-cause mortality

EC Mean (24-h avg): 

0.79

Speciation data from U.S. EPA AQS from 2002

Coefficient (SE) for association with mortality:0.1664 (0.05884)

— Lipfert et al. (2009) 3,065 U.S. 

Counties

All-cause mortality

EC Mean (24-h avg): 

0.82

Estimates from Atmospheric and Environmental Research, Inc (AER) plume-in-grid air quality model

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)

40. Gan et al. (2011) Vancouver, BC CHD 

mortality

BC Mean (24-h avg): 

1.19

High resolution land use regression model

Relative Risk of CHD mortality:1.06 (1.03, 1.09)

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.

Black Carbon Effects on Public Health and the Environment

Report to Congress on Black Carbon 79
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 trafficrelated 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 al. (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-Muñoz 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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Chapter 3﻿

Report to Congress on Black Carbon
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 PM2.5 NAAQS, 

and 2005 ambient PM2.5 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 PM2.5 using a 

global atmospheric model to isolate the total 

anthropogenic contribution to PM2.5 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 PM2.5 were 

estimated to be an order of magnitude higher than 

the impacts of outdoor ozone, due both to high 

PM2.5 concentrations, particularly in very populated 

areas, and a stronger mortality relationship for PM2.5

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 mediumand 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 

PM2.5 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 PM2.5 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.

Black Carbon Effects on Public Health and the Environment

Report to Congress on Black Carbon 81
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 concentrationresponse function for individual pollutants (e.g., 

PM2.5).

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. 

4 http://www.who.int/healthinfo/global_burden_disease/GBD_2005_

study/en/index.html.

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.cira.colostate.edu/improve. For two major PM2.5

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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Report to Congress on Black Carbon
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 (Gomotde 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 

Black Carbon Effects on Public Health and the Environment

Report to Congress on Black Carbon 83
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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Emissions of Black Carbon

4.1 Summary of Key Messages

y 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.

y 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.

y 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.

y 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.

y Long-term historic trends of BC emissions in 

the United States reveal a dramatic increase 

Chapter 4

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.

y 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.

Report to Congress on Black Carbon 85
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 

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 nonopen biomass emissions 

estimates, all BC and 

OC estimates are based 

on EPA’s 2005 modeling 

inventories (termed 

“2005CK” inventories), 

which rely on the 2005 

NEI for PM2.5.

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 

Figure 4-1. BC and OC Fractions of PM2.5 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)

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Report to Congress on Black Carbon
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 

Figure 4-2. Heavy-Duty Diesel PM2.5 Emissions Profile. 

(Source: U.S. EPA, 2002b) 

Emissions of Black Carbon

Report to Congress on Black Carbon 87
(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:

y Open biomass burning (agricultural burning, 

wildfires, and prescribed burning)

y Residential (any combustion for residential 

activities regardless of fuel burned)

y Energy/power (EGUs and other power generation 

sources)

y 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.

Figure 4-3. Contribution to Primary PM2.5, BC, and OC 

Emissions by Mega Source Categories. (Source: U.S. EPA, 

2002a, 2005a)

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Report to Congress on Black Carbon
y Transport/mobile (including on-road, nonroad, 

locomotives, commercial marine, aircraft and tire/

brake wear)

y 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 

Table 4-1. 2005 U.S. Emissions (tons) and Ratios of Emissions by Mega Source Category. (Source: U.S. EPA)

Mega Source Category PM2.5 BC OC OC/BC BC/PM2.5

Open Biomass Burning 2,266,513 224,608 1,058,494 4.7 0.1

Residential 464,063 22,807 204,160 9 0.05

Energy/Power 712,438 43,524 65,138 1.5 0.06

Industrial 219,460 6,085 16,234 2.7 0.03

Mobile Sources (Transport) 626,859 333,400 205,172 0.6 0.53

Other 1,232,123 6,743 112,967 16.8 0.01

Totals (Short Tons) 5,521,456 637,167 1,662,165 2.61 0.12

In GigaGrams (Gg) (5,009) (578) (1,508)

Figure 4-4. U.S. BC Emissions (tons) for Major Source 

Categories. (Source: U.S. EPA)

Emissions of Black Carbon

Report to Congress on Black Carbon 89
Table 4-2. U.S. Emissions of PM2.5, BC, and OC (short tons). (Source: U.S. EPA)

Mega Source 

Category Specific Category

Total 

Primary 

PM2.5

BC Primary OC OC/BC BC/PM2.5

Open Biomass

Wildfires 1,600,358 151,855 738,997 4.9 0.09

Prescribed Burning 535,627 58,525 268,826 4.6 0.11

Agricultural Burning 130,528 14,228 50,671 3.6 0.11

Residential

Residential Wood Combustion 379,878 21,194 200,645 9.5 0.06

Residual Oil Combustion 78,672 787 787 1 0.01

Residential Coal Combustion 2,648 634 1,187 1.9 0.24

Residential Natural Gas Combustion 2,865 192 1,541 8 0.07

Energy/Power

Natural Gas Combustion 64,239 24,668 15,867 0.64 0.38

Bituminous Combustion 394,853 6,697 10,387 1.6 0.02

Sub-Bituminous Combustion 143,383 6,028 4,514 0.7 0.04

Distillate Oil Combustion 23,718 2,372 5,930 2.5 0.1

Wood Fired Boiler 56,289 2,088 19,764 9.5 0.04

Process Gas Combustion 9,457 1,378 2,850 2 0.15

PMSO2 Controlled Lignite 

Combustion 20,499 293 5,826 19.9 0.01

Industrial

Stationary Diesel 4,476 3,452 786 0.2 0.77

Cement Production 17,523 514 2,221 4.3 0.03

Ind Manuf - Avg. 46,501 416 3,422 8.2 0.01

Mineral Products - Avg 23,632 347 1,242 3.6 0.01

Kraft Recovery Furnace 21,222 325 1,111 3.4 0.02

Chem Manuf - Avg 17,526 320 1,608 5 0.02

Lime Kiln 7,002 162 466 2.9 0.02

Heat Treating 14,439 144 1,011 7 0.01

Aluminum Production 5,730 132 223 1.7 0.02

Ferromanganese Furnace 1,240 125 64 0.5 0.1

Surface Coating 9,165 64 1,903 29.7 0.01

Cast Iron Cupola 3,479 32 222 6.9 0.01

Electric Arc Furnace 4,317 16 140 8.8 0

Secondary Aluminum 6,057 12 91 7.6 0

Sintering Furnace 5,739 10 157 15.7 0

Pulp & Paper -Avg 6,569 7 0

Catalytic Cracking 8,864 6 1 0.2 0

Secondary Copper 1,137 1 11 11 0

Ammonium Nitrate Production 1,025 0

Secondary Lead 410 0

Petroleum Ind - Avg 6,224 218 0

Copper Production 432 0

Ammonium Sulfate Production 65 0

Open Hearth Furnace 6,686 1,337 0

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Mega Source 

Category Specific Category

Total 

Primary 

PM2.5

BC Primary OC OC/BC BC/PM2.5

Mobile Sources

On-road diesel 208,473 153,477 44,423 0.3 0.74

Nonroad diesel 145,289 112,058 30,618 0.3 0.77

Locomotive 30,910 22,495 5,130 0.2 0.73

Commercial Marine (C1 & C2) 28,119 21,652 4,937 0.2 0.77

On-road gasoline 75,924 14,510 59,657 4.1 0.19

Nonroad gasoline 55,834 5,444 46,734 8.6 0.1

Commercial Marine (C3) 56,028 1,681 6,303 3.7 0.03

Tire 5,325 1,198 3,060 2.6 0.22

Brakewear 17,801 475 2,321 4.9 0.03

Aircraft 3,156 410 1,988 4.8 0.13

Other

Charbroiling 64,124 2,601 42,975 16.5 0.04

Wood Products - Drying 8,113 649 4,057 6.3 0.08

Paved Road Dust 54,481 569 5,308 9.3 0.01

Dairy Soil 9,862 509 3,139 6.2 0.05

Wood Products-Sawing 12,355 469 5,498 11.7 0.04

Overall Average Manufacturing 10,577 466 927 2 0.04

Unpaved Road Dust 419,648 409 22,897 56 0

Charcoal Manufacturing 5,578 290 100 0.3 0.05

Solid Waste Combustion 14,965 228 1,258 5.5 0.02

Wood Products - Sanding 2,257 135 790 5.9 0.06

Asphalt Manufacturing 2,160 124 93 0.8 0.06

Fiberglass Manufacturing 4,641 93 1,299 14 0.02

Agricultural Soil 334,515 67 10,310 153.9 0

Fly Ash 1,733 30 21 0.7 0.02

Phosphate Manufacturing 992 27 78 2.9 0.03

Industrial Soil 2,011 23 20 0.9 0.01

Food & Ag - Handling 10,331 18 418 23.2 0

Urea Fertilizer 589 12 183 15.3 0.02

Potato Deep-Frying 192 8 121 15.1 0.04

Glass Furnace 7,803 5 55 11 0

Calcium Carbide Furnace 314 4 23 5.8 0.01

Sludge Combustion 163 2 14 7 0.02

Crustal Material 1,160 2 62 31 0

Brick Grinding and Screening 1,272 1 31 31 0

Auto Body Shredding 129 1 10 10 0.01

Inorganic Fertilizer 78 1 2 2 0.01

Asphalt Roofing 1,872 0 1,129 0

Limestone Dust 1,912 0

Sand & Gravel 134,885 0

Construction Dust 96,669 4,463 0

Meat Frying 12,216 7,012 0

Lead Production 33 0

Emissions of Black Carbon

Report to Congress on Black Carbon 91
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 Figure 4-5. U.S. BC Emissions from all Mobile Source 

Categories (333,400 tons). (Source: U.S. EPA)

Mega Source 

Category Specific Category

Total 

Primary 

PM2.5

BC Primary OC OC/BC BC/PM2.5

Other

Synthetic Residential Wood 

Combustion 345 0

Sandblasting 1,673 8 0

Steel Desulfurization 259 0

Inorganic Chemical Manufacturing 4,161 0

Gypsum Manufacturing 1,395 0

Food & Ag-Drying 5,551 666 0

Boric Acid Manufacturing 11 0

Coke Calcining 811 0

Sea Salt 287 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 minor 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.

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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 nonwildfire 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 

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.

and studies looking at differing health impacts of 

PM components including some health studies 

on source apportionment (Thurston et al., 2005; 

Marmur et al., 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).

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 emit about 27,000 tons of BC combined (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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Report to Congress on Black Carbon 93
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; 

Moosmüller et al., 2009). Exactly how much of the 

inventoried OC is light-absorbing is not known at 

this time. 

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 

Figure 4-6. U.S. BC Emissions from all Biomass 

Combustion Source Categories (250,000 short tons). 

(Source: U.S. EPA)

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)

General 

Category Specific Source Total Primary PM2.5

BC OC OC/BC BC/PM2.5

Mega Source 

Category in 

Table 4-2

Open Biomass 

Burning

Agricultural Burning 130,528 14,228 50,671 4 0.11

Open Biomass

Wildfires 1,600,358 151,855 738,997 5 0.09

Prescribed Burning 535,627 58,525 268,826 5 0.11

Subtotal 2,266,513 224,608 1,058,494 4.7 0.1

Residential 

Heating/

Cooking

Residential Wood 

Combustion 379,878 21,194 200,645 9.5 0.06

Residential

Subtotal 379,878 21,194 200,645 9.5 0.06

Other

Wood Fired Boiler 56,289 2,088 19,764 9 0.04 Energy/Power

Charbroiling 64,124 2,601 42,975 17 0.04

Other

Subtotal 120,413 4,689 62,739 13.4 0.04

Biomass Consumption Total 2,766,804 250,491 1,321,878 5 0.09

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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 PM2.5

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-todate 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 

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)

Emissions of Black Carbon

Report to Congress on Black Carbon 95
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 ongoing. 

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.

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 

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 emissions. 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.

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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.

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 

Figure 4-9. BC Emissions by Selected World Region, 2000 (Gg). 

(Source: Lamarque et al., 2010)

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 4-8. Global BC Emissions based on Year 2000 Estimates, in 

Gigagrams (Gg). (Courtesy of Tami Bond, produced based on data from 

Bond et al., 2007)

Emissions of Black Carbon

Report to Congress on Black Carbon 97
Table 4-4. Global BC Emissions in 2000 (in Gg). Transport Includes Aircraft and Shipping. (Source: Lamarque et al., 2010)

Region Energy Industry Transport Residential / Domestic Ag Waste Burning Waste Grasssland Fires Forest Fires Total Country/US Ratio

China 12 669 72 539 44 7 5 9 1,358 3.48Western Africa 0 20 15 127 8 3 505 105 784 2.01India 4 108 74 324 4 2 5 15 538 1.38Brazil 1 53 91 30 3 2 70 215 465 1.19Southern Africa (Rest of) 0 8 5 68 2 0 373 4 460 1.18Indonesia 1 28 34 73 12 1 7 252 407 1.04USA 3 85 216 55 6 3 9 13 390 1Russia 5 33 32 102 7 1 35 145 360 0.92Eastern Africa 0 5 7 119 4 1 210 7 353 0.9Southeastern Asia (Rest of) 1 30 45 101 3 1 6 166 353 0.9South America (Rest of) 0 20 34 30 5 1 42 45 178 0.46

Australia 0 11 12 7 4 0 120 20 174 0.45Western Europe (Rest of) 1 36 88 17 1 1 6 1 150 0.39Central Europe (Rest of) 3 26 40 54 2 1 2 3 131 0.34Japan 2 49 61 7 1 1 0 1 123 0.32South Asia (Rest of) 0 13 30 68 0 1 1 2 116 0.3Middle East 3 37 62 2 6 1 0 0 111 0.29South Korea (Republic of Korea) 3 55 36 9 3 1 0 0 106 0.27

Mexico 3 13 36 6 5 1 8 28 99 0.26

Northern Africa 0 11 36 37 1 1 0 0 87 0.22Central America 0 15 16 12 1 1 2 35 84 0.21Thailand 0 20 33 12 2 0 3 12 83 0.21Canada 0 17 19 4 2 0 5 31 78 0.2France 1 10 48 11 0 0 0 0 71 0.18Ukraine 0 14 9 40 5 0 1 1 71 0.18Argentina 0 12 26 6 7 0 14 4 70 0.18Germany 1 13 48 5 0 1 0 0 68 0.17Asia-"Stan" 0 10 2 27 2 0 25 0 67 0.17South Africa 1 10 14 16 1 0 16 0 58 0.15United Kingdom 1 10 31 4 0 0 0 0 46 0.12

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Region Energy Industry Transport Residential / Domestic Ag Waste Burning Waste Grasssland Fires Forest Fires Total Country/US Ratio

Italy 2 9 31 2 0 0 1 0 46 0.12Taiwan 1 18 12 2 0 0 0 0 32 0.08Venezuela 0 5 7 0 1 0 8 9 30 0.08

Turkey 1 12 10 2 4 0 0 0 30 0.08North Korea (Democratic Peoples Republic) 0 11 1 16 0 0 0 1 29 0.07Baltic States (Estonia, Latvia) 0 1 3 11 0 0 0 1 15 0.04New Zealand 0 1 3 1 0 0 0 0 6 0.01

World Total 54 1,497 1,340 1,947 146 35 1,481 1,128 7,628

Emissions of Black Carbon

Report to Congress on Black Carbon 99
Table 4-5. Global OC Emissions in 2000 (in Gg). Transport Includes Only Aircraft. Global OC Emissions Estimates Not Available for Shipping. 

(Source: Lamarque et al., 2010 )

Region Energy Industry Transport Residential/ Domestic Ag Waste Burning Waste Grasssland Fires Forest Fires Total Country/US Ratio

Western Africa 1 104 43 538 41 3 3,679 882 5,291 6.37Indonesia 5 34 63 327 57 1 51 3,060 3,595 4.33Russia 25 23 33 550 34 3 338 2,582 3,588 4.32China 39 877 72 1,812 208 7 37 122 3,174 3.82Southern Africa (Rest of) 0 24 9 275 7 0 2,732 34 3,083 3.71Brazil 8 203 103 85 14 2 487 1,788 2,690 3.24Eastern Africa 0 22 10 525 20 1 1,461 56 2,095 2.52Southeastern Asia (Rest of) 8 70 80 428 15 1 41 1,405 2,048 2.46India 15 260 63 1,301 20 2 38 146 1,846 2.22Australia 3 8 7 27 19 0 836 165 1,066 1.28South America (Rest of) 4 60 54 116 26 1 312 392 966 1.16USA 72 60 143 198 28 5 97 227 831 1Canada 7 13 14 19 8 0 56 551 669 0.81Mexico 7 20 107 39 22 1 52 265 513 0.62Central America 3 49 29 62 7 1 19 294 463 0.56South Asia (Rest of) 1 45 21 315 1 1 9 27 420 0.5Central Europe (Rest of) 9 19 25 250 10 3 15 49 380 0.46Asia-”Stan” 2 6 5 157 11 2 179 3 364 0.44Ukraine 2 13 7 224 22 1 6 21 297 0.36Thailand 4 51 38 38 8 0 19 102 261 0.31Argentina 2 28 23 8 31 0 103 53 249 0.3Middle East 27 14 171 5 28 1 2 0 248 0.3Western Europe (Rest of) 23 29 36 85 5 2 40 10 230 0.28South Africa 4 17 46 38 6 0 110 0 222 0.27Venezuela 5 6 40 2 3 0 68 76 199 0.24Northern Africa 6 10 51 104 5 1 1 0 179 0.22South Korea (Republic of Korea) 15 71 24 11 13 1 0 4 137 0.17Japan 19 36 30 14 6 1 0 23 130 0.16France 8 6 18 53 0 1 2 4 92 0.11

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Region Energy Industry Transport Residential/ Domestic Ag Waste Burning Waste Grasssland Fires Forest Fires Total Country/US Ratio

Baltic States (Estonia, Latvia, 1 1 2 64 0 0 1 11 79 0.1Turkey 8 9 19 15 17 0 2 0 70 0.08North Korea (Democratic Peoples Republic) 0 13 1 28 0 0 0 16 58 0.07Germany 11 8 21 13 0 1 1 0 56 0.07Taiwan 5 29 12 1 0 0 0 1 46 0.06Italy 9 5 13 8 0 1 5 3 45 0.05United Kingdom 8 6 14 7 0 1 1 0 36 0.04New Zealand 0 1 1 4 1 0 0 0 8 0.01World Total 368 2,249 1,447 7,746 696 47 10,800 12,372 35,725

Emissions of Black Carbon

Report to Congress on Black Carbon 101
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.

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 

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., 

Figure 4-10. Global Distribution of BC and OC Emissions 

by Major Source Category. (Source: Lamarque et al., 2010)

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Table 4-6. OC/BC Ratios by Broad Source Categories. (Source: Lamarque et al., 2010)

Source Category BC (Gg) OC (Gg) OC/BC

Energy/Power 54 368 7

Industry 1,497 2,249 2

Transport/Mobile 1,340 1,447 1

Domestic/Residential 1,947 7,746 4

Open Biomass Burning 2,755 23,868 9

Waste 35 47 1

Totals 7,628 35,725 4.7

Figure 4-11. BC Emissions by World Region, 2000 (Gg). (Source: Lamarque et al., 2010, based on data from 

http://acd.ucar.edu/~lamar/ipcc_ar5/bc.tar.gz) 

Emissions of Black Carbon

Report to Congress on Black Carbon 103
“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 

Figure 4-12. Global BC Emissions by Source Categories and Region. (Source: Lamarque et al., 2010)

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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 BCrelated 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 

Figure 4-13. Geographical Distribution of Global BC Emissions by Latitude. (Source: Bond, 2008)

Emissions of Black Carbon

Report to Congress on Black Carbon 105
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 “topdown” 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 

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 250,499 128,501 51% 49%

Fossil Fuel Combustion 43,049 2,794 6% 1%

Fugitive Dust Sources 1,609 483 30% 0%

Industrial Sources 6,085 1,574 26% 1%

Mobile Sources 333,405 125,784 38% 48%

Other Minor Sources 2,525 755 30% 0%

Totals/averages 637,172 259,891 41%

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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 

Figure 4-14. Comparison of Regional Inventories for China, India, and Indonesia with AR5 Estimates. 

(Source: U.S. EPA)

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Report to Congress on Black Carbon 107
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 regionspecific 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) for a “Waste” category is provided 

in the global inventories, no such estimate was included in the U.S. 

EPA derived inventory. 

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 EPA Source 

Description BC EPA Estimates High By

Energy 3 Energy/Power 39 1200%

Industry 85 Industrial 6 -93%

Transport 216 Mobile Sources 302 40%

Residential 55 Residential 21 -62%

Agricultural Waste Burning 6 Agricultural Burns 13 117%

Waste 3

Grassland Fires 9 Prescribed Burns 53 489%

Forest Fires 13 Wildfires 138 962%

Totals: 390 572 47%

OC Emissions (Gg) in AR5 and EPA Inventories

AR5 Source Description OC EPA Source 

Description OC EPA Estimates High By

Energy 72 Energy/Power 59 -18%

Industry 60 Industrial 15 -75%

Transport 143 Mobile Sources 186 30%

Residential 198 Residential 185 -6%

Agricultural Waste Burning 28 Agricultural Burns 46 64%

Waste 5

Grassland Fires 97 Prescribed Burns 244 151%

Forest Fires 227 Wildfires 670 195%

Totals: 830 1,405 69%

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Report to Congress on 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 “sourcereceptor” 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 midlatitudes 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 longrange 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 

Emissions of Black Carbon

Report to Congress on Black Carbon 109
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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Report to Congress on 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 

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 

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)

Emissions of Black Carbon

Report to Congress on Black Carbon 111
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 PM2.5. 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 PM2.5 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 PM2.5

emissions from industrial sources are estimated to 

have declined only 6% during the entire 1990-2008 

period (U.S. EPA, 2010i).

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 

16 Novakov et al. looked at BC from fossil fuel combustion only.

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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)

Figure 4-18. Historical Reconstruction of Global Emissions Trends. Comparison of Bond et al. (2007) 

(“this work”) with Previous Studies.

Emissions of Black Carbon

Report to Congress on Black Carbon 113
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).

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.

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Observational Data for Black 

Carbon

5.1 Summary of Key Messages

y 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 methoddefined 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 BC:EC 

ratios derived by commercial instrumentation are 

generally within 30%. 

y 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. 

y Ground-level BC measurements across the globe 

indicate estimated concentrations ranging from 

< 0.1 μg/m3 in remote locations to ~15 μg/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 μg/m3 in North America and Europe 

compared to an urban increment of ~6-11 μg/m3

in China. 

y In the United States, BC comprises ~5-10% of 

average urban PM2.5 mass.

Chapter 5

y 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.

y Over the past two decades when U.S. groundlevel 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.

y 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
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.

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”.

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 (Cadle 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 thermaloptical methods that are predominantly used in 

monitoring networks in the United States are the 

thermal-optical transmittance (TOT) and thermaloptical 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 

Table 5-1. Description of BC Measurement Techniques. 

Method Type Method Description Prevalence 

of Use

Light absorption/optical

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. High

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.

Low

Incandescence: Incandescent (glowing) particle mass is measured; BC is quantified by 

calibrating the incandescent signal to laboratory-generated soot. 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.

High

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. Low

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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 

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 BC and BrC.

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)

Observational Data for Black Carbon

Report to Congress on Black Carbon 117
(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 Gelencsér, 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 BC:EC ratios (~0.5) and very high BC:EC 

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. 

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.

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 thermaloptical 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.cira.colostate.edu/improve), 

3

 These BC measurements are publicly available or have been 

included in peer-reviewed publications.

4

 The VIEWS web site (http://vista.cira.colostate.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.

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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 

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)

Observational Data for Black Carbon

Report to Congress on Black Carbon 119
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, 9 The 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 logarithmic 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).

Table 5-2. Summary of Selected Global BC Ambient Concentrations for Urban and Rural/Remote Areas.

Range of Annual Average Concentrations (μg/m3)

Region Networks Year Method Type Urban 

(# Sites)

Rural/Remote

(# Sites)

United States

CSNa/ IMPROVEb

SLAMSc

2005–2007

2007

T

LA

0.3 to 2.5 (~200 sites)

0.3 to 3.0 (~ 45 sites)

0.1 – 0.6 (~150 sites)

Canada NAPSd 2003–2009 T 0.9 – 1.8 (12 sites) 0.4 – 0.8 (4 sites)

Europe EMEPe 2002–2003 T 1.4 – 1.8 (2 sites) 0.2 – 1.8 (12 sites)

Europe EUSAARf 2006 T

LA

1.5 (2 sites)

2.7 (1 site)

0.1 – 0.7 (4 sites)

0.2 – 0.5 (4 sites)

United Kingdom BC Networkg 2009 LA 1.0 – 2.9 (19 sites)

China CAWNETh 2006 T 9.3 – 14.2 (5 sites) 0.3 – 5.3 (13 sites)

Nepal NCO-Pi 2006–2008 LA 0.16 (1 site)

Global 

Background

NOAA GMD Sitesj

Mauna Loa

Point Barrow

South Pole

1990–2006

1988–2007

1987–1990

LA

LA

LA

0.01 – 0.02

0.02 – 0.07

0.002 – 0.004

Other Arctic 

Sites

Alert (Canada) 

Zeppelinfjellf

(Svalbard, Norway)

1989–2008

2002–2009

LA

LA

0.04 – 0.1

0.02 – 0.06

United Kingdom Black Smoke (BS)k 2006 LA 5.0 – 16.0 (12 sites)

a CSN – Primarily urban network sites.

b IMPROVE – Rural network sites.

c BC data at State and Local Air Monitoring Stations from AQS, mostly with Magee Aethalometers.

d Personal communication with Tom Dann (Environment Canada).

e Monitoring was for the period 07/02 – 06/03 from Yttri et al. (2007).

f 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.

g Urban network sites from Butterfield (2010); Curbside site at London Marylebone Road reported 10μg/m3.

h Data and urban/regional/remote classification was for the period 2006 from Zhang et al. (2008).

i Monitoring was for the period 03/06 – 02/08 from Marinoni et al. (2010).

j NOAA Global Monitoring Division Sites – For this table, we modified reported numbers in absorption units using a 

nominal mass extinction coefficient of 10m2g-1. One year from each site was eliminated as non-representative.

k 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 ~40μg/m3 for each year.

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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 μg/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 

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)

Observational Data for Black Carbon

Report to Congress on Black Carbon 121
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 

Figure 5-4. Annual Mean BC Concentrations (μg m-3) 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.cira.colostate.edu/improve/)

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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 PM2.5 (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 

11 Based on 150 filter-based portable samplers and optical 

absorption measurements with the smoke stain reflectometer.

apparent BC concentrations. Actual gradients may 

even be larger. Furthermore, the illustrative withinurban 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 μg/m3 in St. Louis to 

1.2 μg/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 PM2.5 speciation components on an annual and 

seasonal basis can be found elsewhere (Hand et al., 

2011; U.S. EPA, 2009b).

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-toOC 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.

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/)

Observational Data for Black Carbon

Report to Congress on Black Carbon 123
Figure 5-7a. Trends in Black 

Smoke Measurements 

(μg/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 under current 

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 

et al., 2009)

Figure 5-7b. Comparison 

of Ambient Black Smoke 

Measurements (μg/m3, 

annual average) with 

Estimated BC emissions 

(Tg) in the United Kingdom, 

1955-2000. (Source: U.S. EPA)

Figure 5-6. Composition of PM2.5

for 15 Selected Urban Areas in the 

United States. Annual average PM2.5

concentrations (μg/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) 

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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 

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.

Observational Data for Black Carbon

Report to Congress on Black Carbon 125
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 (COH) measurements, 

which are shown to be highly correlated with optical BC. See 

Appendix 1 for further details regarding COHs.

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 et al., 2011)

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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 EPA transitioned 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.

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. 

(Kirchstetter et al., 2008)

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)

Observational Data for Black Carbon

Report to Congress on Black Carbon 127
seasonal variability in BC. However, the higher 

winter BC concentrations, which may be due to local 

emissions patterns or climatology, were not evident 

at all sites and geographic regions.

Based on the evidence provided above from a variety 

of recent BC indicator measurements (and derived 

from different data sources), it appears that ambient 

concentrations of neighborhood/urban and regional 

scale concentrations of BC in the United States 

substantially declined from the mid 1980s. The urban 

and regional BC declines can generally be attributed 

to emissions reductions associated with engine and 

fuel improvements (Bahadur et al., 2011; Kirchstetter 

et al., 2008; Minoura et al., 2006; Murphy et al., 

2011) and reductions in residential biomass burning 

emissions (Burnet et al., 1988; Butler, 1988; Hough, 

1988) to attain NAAQS (Bachmann, 2007) for PM and 

other pollutants. 

5.4.2 Trends in Ambient BC Concentrations 

in the Arctic 

Trends in BC concentrations at three Arctic locations 

(Alert Canada, Barrow Alaska, and Zeppelin Norway) 

are presented in Figure 5-13 (Hirdman et al., 

2010). As stated by the authors, there is a general 

downward trend in the measured BC concentrations 

at all stations, with an annual decrease of −2.1±0.4 

ng/m3 per year (for the years 1989-2008) at Alert 

and −1.4±0.8 ng/m3 per year (2002-2009) at 

Zeppelin. The decrease at Barrow is not statistically 

significant. Based on transport analysis the authors 

conclude that northern Eurasia (the NE - Northeast, 

WNE - West Northeast, and ENE - East Northeast 

clusters) is the dominant emissions source region 

for BC and decreasing emissions in this region drive 

the downward trends. However, there are indications 

that the BC emissions from ENE in wintertime have 

increased over the last decade, probably reflecting 

emissions increases in China and other East Asian 

countries. Emissions associated with the other 

Figure 5-12. Ambient BC Trends (2002-2010), based on monthly distribution of average concentration 

among 15 CSN monitoring locations in the United States. The map shows the location of the 15 

monitoring sites. Starting May 2007, the measured EC were produced with the IMPROVE monitoring 

protocol. The data prior to May 2007 were produced with the EPA NIOSH-type protocol and adjusted to 

represent the IMPROVE protocol. For a more detailed explanation of this graphic, see Appendix A, Figures 

A1-6 and A1-7 and surrounding text. (Source: U.S. EPA)

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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 (Falke et al., 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 

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)

Observational Data for Black Carbon

Report to Congress on Black Carbon 129
Figure 5-15. Absorption Angstrom Exponent (AAE) Values for AAOD Spectra Derived from AERONET 

Data. Black: Urban/Industrial or Mixed; Green: Biomass Burning; Red: Desert Dust. Shading for each 

location indicates wavelength pair (in nm) for AAE calculation. GSFC=Goddard Space Flight Center, 

Greenbelt, MD. (Russell et al., 2010)

Figure 5-14. Aerosol Absorption Optical Depth (AAOD) from AERONET (1996-2006) and OMI (2005-

2007). (a) Aerosol absorption optical depth, AAOD, (x100) 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)

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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. 

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)

Observational Data for Black Carbon

Report to Congress on Black Carbon 131
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 snowcovered 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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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 

et al. (2006)) 

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)

Observational Data for Black Carbon

Report to Congress on Black Carbon 133
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; Flanner 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 

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 Cachier and Pertuisot (1994), Chylek et al. (1987; 1992; 1995), Ross Edwards (personal 

communication regarding WAIS Divide ice core), Kaspari et al. (2011), Lavanchy et al. (1999), Legrand 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))

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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 longdistance 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 

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 et al., 2003)

Observational Data for Black Carbon

Report to Congress on Black Carbon 135
(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.

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 monthlyresolution 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), nonsea-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. 

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))

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Report to Congress on 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, 

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 et al., 2010b).

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)

Observational Data for Black Carbon

Report to Congress on Black Carbon 137

Benefits of Reducing Black Carbon 

Emissions

6.1 Summary of Key Messages

y 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.

y 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 

PM2.5 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. 

Chapter 6

– Programs to reduce PM2.5 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 PM2.5-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. 

y 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 longterm 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.

Report to Congress on Black Carbon 139
– 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.

y 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.

y 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 welldeveloped, 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 coemitted 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 EPA’s 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 PM2.5 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 PM2.5 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). EPA’s analyses of 

these benefits provide useful information regarding 

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Chapter 6﻿

Report to Congress on Black Carbon
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)

Quantified and Monetized PM2.5 Health Endpoints Un-Quantified and Non-Monetized 

PM2.5 Health Endpoints

Premature mortality based on cohort study estimates and expert 

elicitation estimates

Low birth weight, pre-term birth and other reproductive 

outcomes

Hospital admissions: respiratory and cardiovascular Pulmonary function

Emergency room visits for asthma Chronic respiratory diseases other than chronic bronchitis

Nonfatal heart attacks (myocardial infarctions) Non-asthma respiratory emergency room visits

Lower and upper respiratory illness UVb exposure (+/-)

Minor restricted activity days

Work loss days

Asthma exacerbations (among asthmatic populations)

Respiratory symptoms (among asthmatic populations)

Infant mortality

Benefits of Reducing Black Carbon Emissions

Report to Congress on Black Carbon 141
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 6-3. 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 Benefitsb Annual Costs Benefit/Cost 

Ratio Benefit Year

Transportation

 Light Duty Tier 2 $34 $7.2 4.7 2030

 Heavy Duty 2007 $92 $5.5 16.7 2030

 Non-road Diesel Tier 4 $105 $2.5 41.5 2030

 Locomotive & Marine Diesel $9.9–23.8 $0.8 13.1–31.4 2030

 Ocean Vessel Strategy $119–292 $3.4 35.5–87.1 2030

Stationary Sources

 2006 PM NAAQSc $22 $7.1 3.1 2020

 Cement NESHAP $8.3–20.1 $1.01–1.04 8.2–19.4 2013

 Stationary Spark Ignition RICE NESHAP $0.52–1.2 $0.26 2–4.7 2013

 Stationary Compression Ignition Engine NESHAP $0.95–2.3 $0.37 2.5–6.2 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.

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 (2011a) Table 5-6, adjusted from 2006$ to 2010$)

Health Effect Reductions

Year 2010 Year 2020

Incidence Avoided Valuation 

(millions 2010$)

Incidence 

Avoided

Valuation 

(millions 2010$)

Adult Mortality 160,000 $1,300,000 230,000 $1,800,000

Infant Mortality 230 $2,100 280 $2,700

Chronic Bronchitis 54,000 $26,000 75,000 $39,000

Acute Bronchitis 130,000 $66 180,000 $100

Acute Myocardial Infarction 130,000 $15,000 200,000 $23,000

Hospital Admissions, Cardiovascular 45,000 $1,400 69,000 $2,200

Lower Respiratory Symptoms 1,700,000 $32 2,300,000 $45

Upper Respiratory Symptoms 1,400,000 $45 2,000,000 $65

Asthma Exacerbation 1,700,000 $97 2,400,000 $140

Lost Work Days 13,000,000 $2,200 17,000,000 $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.

142

Chapter 6﻿

Report to Congress on Black Carbon
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 copollutants. 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 et al. (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 PM2.5 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 benefitper-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.

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)

Source Category Monetized Benefit 

Per Ton in 2015

Monetized Benefit 

Per Ton in 2020

Monetized Benefit 

Per Ton in 2030

Area Source

 Pope et al. (2002) $360 $400 $480

 Laden et al. (2006) $880 $970 $1,200

Mobile Source

 Pope et al. (2002) $280 $300 $360

 Laden et al. (2006) $680 $740 $890

EGU and Non-EGU

 Pope et al. (2002) $230 $250 $290

 Laden et al. (2006) $560 $610 $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.

Benefits of Reducing Black Carbon Emissions

Report to Congress on Black Carbon 143
Fann, et al. (2009) that used an innovative reducedform 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.html.

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 extraregional 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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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:

y Use of diesel particle filters as part of a Euro VI 

package for on-road and off-road diesel vehicles

Abbreviation Region

NA North America

SA South America

EU Europe

FSU Former Soviet Union

AF/ME Africa/Middle East

IN South Asia (India)

EA East Asia (China)

SE/AU 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 et al., 2011)

Note:

Confidence intervals (95%) reflect uncertainty in the Concentration Response Function only.

Benefits of Reducing Black Carbon Emissions

Report to Congress on Black Carbon 145
y Introduction of clean-burning cook stoves for 

cooking and heating in developing countries

y Replacement of traditional brick kilns with vertical 

shaft kilns and Hoffman kilns 

y Elimination of high-emitting vehicles in on-road 

and off-road transport (excluding shipping)

y Ban of open field burning of agricultural waste

y 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 

(Shindell et al., 2012; UNEP 

and WMO, 2011b). Consistent 

with the results of Anenberg 

et al. (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 

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. 

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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 PM2.5 constituents, 

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 

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, 2011b)

Benefits of Reducing Black Carbon Emissions

Report to Congress on Black Carbon 147
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 highemitting 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 

PM2.5-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, PM2.5-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 coemitted 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 2oC 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 coemitted 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 broadscale 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 coemitted 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) 

Benefits of Reducing Black Carbon Emissions

Report to Congress on Black Carbon 149
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 m-2), 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 CH4 emission 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).

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 Unger et 

al. (2010) (also shown in Chapter 2).

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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 mediumduty 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 

Figure 6-5. Observed Deviation of Temperature to 2009 and Projections under Various Scenarios. 

Immediate implementation of the identified BC and CH4 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. 

(UNEP and WMO, 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.

Benefits of Reducing Black Carbon Emissions

Report to Congress on Black Carbon 151
of 1.4 W m-2 (±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 IceCovered 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:

[I]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 

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 of 40°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. ...

[E]missions 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 

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 icecovered 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 

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)

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Report to Congress on Black Carbon 153
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 snowand 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: 

y 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). 

y 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. 

y Effects of non-BC aerosols in the Arctic. 

y 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 longterm 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 longlived 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 2oC since 

preindustrial times was adopted as an international target under 

the UN’s Copenhagen Accord in December 2009. 

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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 coemissions 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-ofpipe 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 coemissions 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 

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)

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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 

by BC. 

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 (IAMs) 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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damages. However, there are a number of factors 

that would complicate these calculations. These 

difficulties stem partly from limitations in the 

capabilities of IAMs, 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 IAMs 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 IAMs 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 IAMs may not be preferred. Rather, a bottomup 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 

Figure 6-9. Cause and Effect Chains for (a) CO2 and (b) BC from Emissions to Damages. (Source: U.S. EPA)

(a)

(b)

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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 costeffectiveness 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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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 PM2.5 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. PM2.5 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 PM2.5 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 PM2.5 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 PM2.5 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 PM2.5, 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 colocated 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 nearterm. 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 BCemitting 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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Mitigation Overview: Designing 

Strategies for Public Health and 

Near-Term Climate Protection

7.1 Summary of Key Messages

y 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.

y 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.

y Available control technologies can provide lowcost 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.

Chapter 7

– 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. 

y 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). 

y 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.

y 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

y Considering the location and timing of emissions 

and accounting for co-emissions will improve 

Report to Congress on Black Carbon 161
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 snowcovered regions. This would include BC 

emitted directly in those areas as well as BC 

transported into those areas from other areas.

y 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. 

y 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 decisionmakers 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 NOx) 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 (2010i) shows that U.S. emissions of direct 

PM2.5 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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Report to Congress on Black Carbon 163
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 by Jacobson 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 m-2. 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 

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)

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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)

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 – Riahi et al. (2007)

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Report to Congress on Black Carbon 165
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 

1 http://www.iiasa.ac.at/web-apps/tnt/RcpDb/dsd?Action=htmlpage&

page=welcome#

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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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 decisionmakers 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 

Figure 7-4. Policy Framework for Black Carbon Mitigation Decisions. (Source: U.S. EPA)

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Report to Congress on Black Carbon 167
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:

y 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. 

y 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-ofpipe 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:

y 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.

y 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. 

y 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.

y 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 

Mitigation Overview: Designing Strategies for Public Health and Near‐Term Climate Protection

Report to Congress on Black Carbon 169
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 number of source categories. 

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)

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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.

y 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. 

y 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.

y 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 

Mitigation Overview: Designing Strategies for Public Health and Near‐Term Climate Protection

Report to Congress on Black Carbon 171
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.

y 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 al., 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 (multipollutant, 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 multipollutant assessment can produce more costeffective 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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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 impact 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-HimalayanTibetan 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, 

Mitigation Overview: Designing Strategies for Public Health and Near‐Term Climate Protection

Report to Congress on Black Carbon 173
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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Mitigation Approaches for Mobile 

Sources

8.1 Summary of Key Messages

y 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.

y 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.

y 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.

Chapter 8

y 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.

y 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 

Report to Congress on Black Carbon 175
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 lowsulfur 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 lowsulfur 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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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 [NOx]) 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 emissions 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.

Mitigation Approaches for Mobile Sources

Report to Congress on Black Carbon 177
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 

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 1990 2005 2020 2030 19902005 20052030

Onroad gasoline 69,629 14,510 9,538 10,027 -79% -31%

Onroad diesel 219,958 153,477 28,175 7,615 -30% -95%

Tire 809 1,198 1,435 1,720 48% 44%

Brakewear 290 475 569 682 64% 44%

Nonroad gasoline 5,420 5,444 4,702 5,174 0% -5%

Nonroad diesel 148,537 112,058 31,254 9,356 -25% -92%

Commercial Marine (C1 & C2) 22,122 21,652 11,595 5,440 -2% -75%

Commercial Marine (C3) 1,262 1,681 864 1,306 33% -22%

Locomotive 19,317 22,495 11,349 5,684 16% -75%

Aircrafta 283 410 457 553 45% 35%

Total BC Emissions (Mobile) 487,628 333,400 99,940 47,557 -32% -86%

ORGANIC CARBON

Onroad gasoline 262,065 59,657 43,711 47,421 -77% -21%

Onroad diesel 66,056 44,423 14,883 10,580 -33% -76%

Tire 1,734 3,060 3,678 4,407 76% 44%

Brakewear 1,191 2,321 2,790 3,343 95% 44%

Nonroad gasoline 37,613 46,734 41,137 45,424 24% -3%

Nonroad diesel 33,872 30,618 9,759 3,891 -10% -87%

Commercial Marine (C1 & C2) 5,045 4,937 2,772 1,710 -2% -65%

Commercial Marine (C3) 4,734 6,303 8,644 13,060 33% 107%

Locomotive 4,405 5,130 2,659 1,507 16% -71%

Aircrafta 1,372 1,988 2,217 2,682 45% 35%

Total OC Emissions (Mobile) 418,088 205,172 132,252 134,025 -51% -35%

DIRECT PM2.5

Onroad gasoline 335,205 75,924 54,682 59,106 -77% -22%

Onroad diesel 290,478 208,473 43,698 18,765 -28% -91%

Tire 3,678 5,325 6,450 7,727 45% 45%

Brakewear 11,129 17,801 21,559 25,830 60% 45%

Nonroad gasoline 54,198 55,834 49,000 54,078 3% -3%

Nonroad diesel 192,905 145,289 46,310 18,463 -25% -87%

Commercial Marine (C1 & C2) 28,730 28,119 15,789 9,741 -2% -65%

Commercial Marine (C3) 42,082 56,028 14,407 21,767 33% -61%

Locomotive 25,087 30,910 15,145 8,584 23% -72%

Aircrafta 2,178 3,156 3,519 4,257 45% 35%

Total PM2.5 Emissions (Mobile) 985,671 626,859 270,559 228,318 -36% -64%

a Non landing and take-off (LTO) emissions not included; also, planned technology and operations improvements that 

require funding for implementation are not included in the forecast.

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Report to Congress on Black Carbon
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 

6

 See http://www.epa.gov/otaq/climate/regulations.htm.

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)

Mitigation Approaches for Mobile Sources

Report to Congress on Black Carbon 179
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 lbs. 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 precontrol 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 lbs 

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 emissions 

characterization with four 2007 heavy duty diesel engines. See 

Khalek et al. (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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Chapter 8﻿

Report to Congress on Black Carbon
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 lowexhaust 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). 

Mitigation Approaches for Mobile Sources

Report to Congress on Black Carbon 181
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/highwaydiesel/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).

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)

Rule Estimated Cost (2010$) Per Ton PM2.5 Reduced

NPV, 3% rate NPV, 7% rate

Heavy-Duty Diesel Rule (2001) $16,652 $19,216 

Nonroad Diesel Rule (2004) $13,762 $14,461 

Locomotive/Marine Rule (2008) $8,579 $9,778 

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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) or 4-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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Report to Congress on Black Carbon 183
Commercial marine vessels are classified as C1, C2, 

and C3 based on engine size. C1 marine engines 

are similar in size (less than 5 l/cylinder or for some 

categories less than 7 l/cylinder) to those used in 

construction/farm equipment. C2 marine engines 

(between 5 or 7 and 30 l/cylinder) are similar to 

locomotive diesels. The C3 engines (greater than 

30 l/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 C1 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 pervessel 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 onroad 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 onroad 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 heavyduty 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.

Mitigation Approaches for Mobile Sources

Report to Congress on Black Carbon 185
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, 

14 See http://www.meca.org/cs/root/diesel_retrofit_subsite/what_is_

retrofit/what_is_retrofit.

15 See http://www.epa.gov/cleandiesel/technologies/retrofits.htm.

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:

y 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 

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.nrel.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/100.

Mitigation Approaches for Mobile Sources

Report to Congress on Black Carbon 187
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.

y 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.

y Closed Crankcase Ventilation Systems (CCVS) 

provide negligible BC reductions. In many 

diesel engines, crankcase emissions or “blowby” 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.

y 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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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 

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’s Tier 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.

Mitigation Approaches for Mobile Sources

Report to Congress on Black Carbon 189
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-rollingresistance 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:

y Automatic engine shut-off devices programmed 

to shut down the engine after a preset time limit

y Direct-fired heaters to eliminate idling used to 

heat the cab 

y Auxiliary power units (APU) or generators to 

provide power for cab comfort at rest stops and 

eliminate the need to run the truck engine

y 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 antiidling 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 PM2.5

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 costeffectiveness 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/PLAW109publ58.pdf

Mitigation Approaches for Mobile Sources

Report to Congress on Black Carbon 191
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 cargohauling 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 SmartWaybased 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 heavyduty 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-thanrequired 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 1B 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

20 http://www.tceq.texas.gov/airquality/terp.

Mitigation Approaches for Mobile Sources

Report to Congress on Black Carbon 193
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 wellmaintained 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. Lowsulfur 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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Mitigation Approaches for 

Stationary Sources

9.1 Summary of Key Messages

y 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.

y 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.

y 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.

y 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. 

Chapter 9

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 PM2.5 (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.

Report to Congress on Black Carbon 195
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 nearsteady 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#feb11. 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 PM2.5 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 μm in diameter and larger and 99.6% to 

99.9% for particles 2.5 μm in diameter and smaller. 

Studies have shown that collection efficiencies 

greater than 99% can be achieved for particles less 

than 1 μm 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 woodfired boilers found that mechanical collectors, such 

as multicyclones, were only modestly effective in 

reducing PM emissions, with an average of about 

Mitigation Approaches for Stationary Sources

Report to Congress on Black Carbon 197
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 costeffectiveness 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 

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/fr28jn11.pdf. 

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 PM2.5 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 costeffectiveness 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 

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/regdata/RIAs/boilersriafinal110221_psg.pdf, Tables 7-2 and 

7-3.

Mitigation Approaches for Stationary Sources

Report to Congress on Black Carbon 199
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 tuneup. 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 

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$.)

Fuel Type Technology PM Reduction 

Potential

Size of 

ICI Boiler 

(mmBTU/hr)

Capital 

Costs 

(Million, 

2010$)

Cost Effectiveness ($/

ton removed, 2010$) Reference

Coal Dry ESP 90–99% 250 $3.5 – 47 $184 – 1,529 MACTEC (2005)

Coal Fabric Filter 90–99% 250 $2.4 – 27 $498 – 1,183 MACTEC (2005)

Oil Dry ESP 90–99% 250 $2.4 – 26 $2,739 – 24,715 MACTEC (2005)

Wood ESP 99.5% Medium — $239 – 344 STAPPA (2006)

Wood Fabric Filter 99.5% Medium — $173 – 293 STAPPA (2006)

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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 energyefficient 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 

9

 See http://www.resourceefficientbricks.org/background.php. 

Mitigation Approaches for Stationary Sources

Report to Congress on Black Carbon 201
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 fineparticle 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 (BuzcuGuven 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).

Mitigation Approaches for Stationary Sources

Report to Congress on Black Carbon 203
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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Mitigation Approaches for 

Residential Heating and Cooking

10.1 Summary of Key Messages

y 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 gasburning) 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 $1,800/ton PM2.5 reduced (2010$).

– 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.

Chapter 10

y 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-HimalayanTibetan 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 205
– 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 highquality 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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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 PM2.5 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 PM2.5 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.

Mitigation Approaches for Residential Heating and Cooking

Report to Congress on Black Carbon 207
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.

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 woodfired 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.

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 Woodburning 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 PM2.5

Figure 10-1. OC/BC Emission Ratios by Source Category and Fuel Type. 

(Source: U.S. EPA)

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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-1; 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 lb/hr) tested by at an accredited laboratory; 

and a maximum weight of 800 kg (1,760 lb). Many 

types of sources are exempt from the existing NSPS, 

including:

y Wood heaters used solely for research and 

development purposes

y Wood heaters manufactured for export (partially 

exempt)

y Coal-only heaters

y Open masonry fireplaces constructed on site

y Boilers

y Furnaces

y 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 

Mitigation Approaches for Residential Heating and Cooking

Report to Congress on Black Carbon 209
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. 

Some of the benefits of 

replacing inefficient wood 

stoves include:

y Reduction in PM2.5 and 

toxic air pollutants (e.g., 

benzeno(a)pyrene) by 70%

y Reduction in indoor PM2.5

emissions by 70% according 

to University of Montana3

3

 For more information, see: http://www.ncbi.nlm.nih.gov/

pubmed/18665872.

y Improvement in energy efficiency by 50%, using 

one-third less wood

y 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 PM2.5 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 

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$.

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

y 815 tons of CO2

y 53 tons of CH4

y 27 tons of PM2.5

y 4 tons of toxic pollutants

y 14 tons of OC

y 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.)

Figure 10-2. Cost Per Ton PM2.5 Reduced for Replacing Non-EPACertified Wood Stove with EPA-Certified Woodstove (in 2010$). 

(Source: U.S. EPA, based on data from http://www.marama.org/visibility/

ResWoodCombustion/RWC_FinalReport_121906.pdf)

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Mid-Atlantic Regional Air Management Association 

(MARAMA) document called Control Analysis and 

Documentation for Residential Wood Combustion in 

the MANE-VU Region (2006). This document focused 

on the costs of total PM2.5 mitigation. The results 

suggest that the cost per ton of PM2.5 emissions 

reduced from wood stove changeouts and fireplace 

retrofits is similar to the cost of other PM2.5 controls 

discussed in previous chapters of this report. 

Figures 10-2 and 10-3 summarize the MARAMA 

estimates of the cost per ton PM2.5 reduced from 

these measures; these costs vary depending on the 

type of wood burning appliance being replaced 

(old wood stove vs. open fireplace) and on the 

replacement technology (e.g., EPA-certified wood 

stove vs. wood pellet stove). 

10.3.4.2 Outdoor Wood-Fired Hydronic Heater 

Program

In 2007, EPA initiated a partnership to reduce 

emissions from new outdoor wood-fired hydronic 

heaters. This program is aimed largely at areas 

with PM2.5 air quality problems. EPA has worked 

with industry to reach agreement on voluntary 

performance levels for new heaters to bring them to 

market faster than feasible under regulation. Similar 

to the wood stove changeout program, there are 

potential climate change, air quality, and energy 

efficiency benefits with this program. The program 

is structured in two phases: under Phase 1, qualified 

new units were 70 % cleaner than existing units and, 

under Phase 2, which began in October 2008, new 

units are required to be 90% 

cleaner than existing units.5 EPA 

has now expanded the program 

to include indoor models and 

hydronic heaters that are fueled 

by other kinds of solid biomass 

(e.g., wood pellets).

As of 2011, nearly 10,000 

EPA‐qualified units had been 

sold; 24 manufacturing partners 

had agreed to produce units 

70%-90% cleaner; and 22 

models had been placed on the 

market, reducing approximately 

6,100 tons of PM2.5 emissions 

each year after 2011 and 

providing approximately $2.2 

billion to $5.4 billion (2010$) in 

estimated health benefits in the 

U.S.6

10.3.4.3 New Construction 

Wood-Burning Fireplace 

Program

The EPA voluntary Wood-burning Fireplace Program 

is modeled after the Hydronic Heater Program 

and helps reduce wood smoke emissions growth 

in areas with PM2.5 air-quality problems. The twophase program covers new installation of low mass 

(i.e., pre-manufactured) and masonry fireplaces, 

and is expected to drive technology improvements 

much sooner than possible through regulation. 

The program qualifies models achieving a Phase 1 

(34% reduction) or a Phase 2 (54% reduction) PM2.5

emission level. EPA has worked closely with the 

hearth products industry to develop this program; 

however, growth in the program has been hampered 

by the slowdown in new home construction in the 

United States.

10.3.5 Additional Regulatory Approaches to 

Limiting Wood Smoke Emissions

A variety of regulatory programs, including wood 

burning curtailment programs and requirements 

to remove old stoves upon resale of a home, have 

proven effective in helping to address wood smoke. 

5

 Use of Phase 1 labels was prohibited after March 31, 2010. See 

http://www.epa.gov/burnwise/owhhlist.html.

6

 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$.

Figure 10-3. Cost Per Ton PM2.5 Reduced ($/Ton) for the Addition of an 

Insert into a Fireplace (2010$). (Source: U.S. EPA, data courtesy of MARAMA)

Mitigation Approaches for Residential Heating and Cooking

Report to Congress on Black Carbon 211
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 EPAcertified 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 μg/m3. The cost effectiveness was 

estimated to be approximately $6,300 - $11,100 

per ton of PM2.5 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/ttncaaa1/t1/memoranda/strategiesdoc-8-11-09.pdf.

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

y 3% percent of the respondents said the workshops brought about positive change on how they burned wood

y 34% have updated their wood burning appliances; 90% of those chose EPA approved appliances

y 41% of those surveyed have changed out or intend to change out their old wood burning appliances for cleaner 

technology.

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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 SubSaharan 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 

μg/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 

Mitigation Approaches for Residential Heating and Cooking

Report to Congress on Black Carbon 213
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 SubSaharan 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, individualspecific 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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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 

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).

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:

y electric stoves

y gas and liquid fuels

y processed solid fuels

y advanced biomass stoves

y rocket stoves

Figure 10-4. The Turbococina 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 René Nuñez 

Suárez, Turbococina)

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)

Mitigation Approaches for Residential Heating and Cooking

Report to Congress on Black Carbon 215
y simple stoves

y solar cookers

y 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 

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 

Figure 10-6. CleanCook Ethanol 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)

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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 cleanburning 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 projectbased 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 

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)

(a) (b) (c)

(d) (e)

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Report to Congress on Black Carbon 217
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 realworld 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 BC:OC 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%), PM2.5 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 cookingtime BC concentrations were reduced to 5-100 μg/

m3 by the top-performing forced draft stove (as 

compared to 50-1000 μg/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 BC:OC 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 (~15¢) per kilogram of pellets.

Figure 10-8. Philips Woodstove (forced draft) 

Manufactured in Lesotho. (Photo courtesy of Stephen 

Walker, African Clean Energy Ltd.)

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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 EPA’s 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 (Adkins et al., 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 PM4.0 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 

(7.2%). 

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 

Figure 10-9. Rocket Stoves. (a) Envirofit G3300 Biomass Cookstove in Use in Tanzania. (Photo courtesy 

of Nancy Ryden, Envirofit) (b) Zoom Dura Biomass Cookstove produced by EcoZoom. (Photo courtesy of 

EcoZoom)

(a) (b)

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Report to Congress on Black Carbon 219
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 

metallic plancha (for making tortillas), and durability. 

Certain models of these stoves have combustion 

chambers that might also be used to build qualitycontrolled 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 

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 

Figure 10-10. Prakti Double-Pot Woodstove with Chimney. (Photo 

courtesy of Mouhsine Serrar, Prakti Design)

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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 

Mitigation Approaches for Residential Heating and Cooking

Report to Congress on Black Carbon 221
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:

y 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 Coloradobased EnviroFit International on sales); and major 

corporations building their business in emerging 

markets (e.g., Philips, Bosch-Siemens). 

y 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).

y 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.

y 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 

Figure 10-11. Number of Improved Stoves Sold by PCIA 

Partners, 2003-2010. (Source: U.S. EPA)

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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 wellconstructed monitoring and evaluation programs 

were other tools highlighted as critical to success in 

reaching the poor (World Bank, 2010). Other major 

considerations include:

y 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.

y 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 startup. Separate financing tools are needed make 

advanced stoves affordable for the poorest 

populations.

y 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).

y Global Leadership: Coordination and crossdisciplinary 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 over time. 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 

Mitigation Approaches for Residential Heating and Cooking

Report to Congress on Black Carbon 223
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 

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.)

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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 noncommunicable 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 

some way before they can meet a health-based standard, but they 

are much, much better than the traditional stoves we have now” 

(Adler, 2010).

Mitigation Approaches for Residential Heating and Cooking

Report to Congress on Black Carbon 225

Mitigation Approaches for Open 

Biomass Burning

11.1 Summary of Key Messages

y 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. 

y 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. 

y 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 landChapter 11

management objectives in fire-dependent 

ecosystems.

y 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.

Report to Congress on Black Carbon 227
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-andburn 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 

Type of 

Burning Description Land Type Typical Resource Management Objective(s)

Agricultural

The planned burning of vegetative debris from 

agricultural operations. (Domestic)

Forestland, 

cropland, 

rangeland, 

grassland, 

wetlands

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.

The use of fire as a method of clearing land for 

agricultural use or pastureland. (International)

Forestland, 

rangeland, 

grassland, 

wetlands

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.

Table 11-1. Types of Open Biomass Burning. (U.S. EPA, 1998)

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Report to Congress on Black Carbon
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 snowfed 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, 

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. 

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 longneedle 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) 

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.

Mitigation Approaches for Open Biomass Burning

Report to Congress on Black Carbon 229
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 firedependent 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/

stelprdb1046311.pdf. 

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Report to Congress on Black Carbon
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, sitespecific 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

y 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).

y 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.

Mitigation Approaches for Open Biomass Burning

Report to Congress on Black Carbon 231
– 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).

y Reduce fuel loadings

– Remove straw/stubble before the burn.

y 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. 

y Convert Land Use 

– Convert from a crop that requires burning to a 

crop that does not. 

– Convert land to non-agricultural use.

y Educate Farmers

– Provide training to farmers on proper burning 

techniques that reduce emissions.

11.5.1.2 Prescribed Burning PM Mitigation

y 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).

y 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.

y 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.

6

 See http://www.biochar.org.

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– 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 

rangelands and other lands can reduce grassy 

or brushy fuels prior to burning, and can help 

reduce burn frequency.

y 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

y 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 

7

 See http://www.fs.fed.us/eng/pubs/html/05511303/05511303.html.

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 sitespecific 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). 

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.

Mitigation Approaches for Open Biomass Burning

Report to Congress on Black Carbon 233
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 stateprotected 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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Report to Congress on Black Carbon
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 multiagency program contains the following components 

(USDA, 2010):

y 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 longrange 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 longrange 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).

y 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. 

Mitigation Approaches for Open Biomass Burning

Report to Congress on Black Carbon 235

Key Black Carbon Mitigation 

Opportunities and Areas for Further 

Research

12.1 Summary of Key Messages

y 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.

y 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 bestsuited to particular locations or situations. 

y 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 

Chapter 12

if such controls are more costly than controls 

on secondary PM precursors, the combined 

public health and climate benefits may justify 

the expense.

y 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. 

Report to Congress on Black Carbon 237
– 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 coalfired brick kilns); and transportation, primarily 

on-road and off-road diesel engines.

y 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.

y 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)

y 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 snowand 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 nearterm 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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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-sizefits-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 PM2.5 mass in most 

areas; second, they are the pollutants principally 

responsible for regional transport of PM2.5, and 

therefore have been the focus of national rules; 

third, reductions in NOx and SOx are highly costeffective. Widespread reductions in NOx and SOx

emissions have been credited with substantial 

decreases in ambient PM2.5 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 PM2.5

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 PM2.5 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, 

Key Black Carbon Mitigation Opportunities and Areas for Further Research

Report to Congress on Black Carbon 239
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 PM2.5 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/deqaqd-air-aqe-sip-pm25-1-14-08_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 LibertyClariton PM2.5 Nonattainment Area. Available online at: http://www.

achd.net/airqual/Liberty-Clairton_PM2.5_SIP-Apr2011.pdf.

4

 Alabama State Implementation Plan for PM2.5 for Birmingham, 

Alabama.

less advanced. For example, some megacities in Asia 

have annual average PM2.5 concentrations above 100 

μg/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 PM2.5 reductions 

in developing countries are large, and controls 

on BC and other directly emitted PM2.5 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. 

y 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.

y 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 PM2.5 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.

y 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 

Key Black Carbon Mitigation Opportunities and Areas for Further Research

Report to Congress on Black Carbon 241
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 PM2.5

(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:

y 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. 

y 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 endof-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. 

y 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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Report to Congress on Black Carbon 243
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):

y 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.

y 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 pointof-manufacture performance standards, and 

work to develop new technologies and incentive 

programs to replace old stoves and boilers.

y 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:

y 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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y 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.

y Transportation, primarily on-road and offroad 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 

Key Black Carbon Mitigation Opportunities and Areas for Further Research

Report to Congress on Black Carbon 245
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, 

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. 

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 

Figure 12-1. Key Policy-Relevant Scientific Uncertainties Related to BC. 

(Source: U.S. EPA)

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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; highemitting 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, 

Key Black Carbon Mitigation Opportunities and Areas for Further Research

Report to Congress on Black Carbon 247
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, multiwavelength 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 snowand 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 

sulfate, 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 nonBC 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. 

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Chapter 12﻿

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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. 

Key Black Carbon Mitigation Opportunities and Areas for Further Research

Report to Congress on Black Carbon 249

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 

Appendix 1

in aircraft and on remote sampling platforms. 

Globally, a significant amount of ambient data 

has been compiled from the following types of 

measurements:

y Ground-based ambient air measurements are 

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 A1-1. 

y Ice core measurements of BCa and ECa have been 

conducted in glaciers around the world, providing 

a historical record of BC concentrations. 

y 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 lightabsorbing 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 A1-1). The second major 

class of methods, optical methods, quantifies the 

light absorbing component of particles as BCa, 

Report to Congress on Black Carbon 251
252

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Report to Congress on Black Carbon

Table A1-1. Global Monitoring Activities. (Source: U.S. EPA)

Worldwide Air Monitoring Networksa, b for Black Carbon

Network Country/ Agency Years of Data BC - Indicator Number of Sites Measurement Method Location of Information and/ or Data

ESRL/GMD Aerosol NetworkBaseline StationsRegional StationsMobile/Cooperative Platforms

United States / NOAA

1957–Present BCa and/

or Aerosol optical properties

4 Rural3 Rural15 Rural

Aerosol Monitoring System – Aethalometers, Particle Soot/Absorption Photometers, Nephelometers, etc.

http://www.esrl.noaa.gov/gmd/aero/index.html

World Data Centre for Aerosols Global 

AtmosphericWatch

1974–Present BCa and/or

Aerosol optical properties

~16 Rural Aerosols – Light Absorption/

EBC, AOD, Light scattering & back scattering, Size distribution

http://wdca.jrc.it/http://gaw.tropos.de/gaw_program.html

Nepal Climate Observatory-Pyramid (NCO-P)

Nepal 2006–2008 BCa 1 Rural MAAP http://www.atmos-chem-physdiscuss.net/10/8379/2010/acpd-10-

8379-2010.pdf 

CSN/STN—PM2.5 Speciation Trends Networkc

United States / EPA

1999–Present ECa ~200 urban Thermal Optical Transmittance http://www.epa.gov/ttnamti1/

specgen.html

IMPROVE—Interagency Monitoring of Protected Visual Environments

United States / NPS

1988–Present ECa 110 rural 

(plus ~67 protocol sites)

Thermal Optical Reflectance http://vista.cira.colostate.edu/

IMPROVE/

ARIES / SEARCH—Aerosol Research Inhalation Epidemiology Study / SouthEastern Aerosol Research and Characterization Study experiment

United States /EPRI / SC

1992–Present ECa 5 Urban

3 Rural

Thermal Optical Reflectance http://www.atmospheric-research.

com/studies/SEARCH/index.html

NAPS—National Air Pollution Surveillance Network

Canada 2003–Present ECa 4 rural

13 urban

R&P Partisol-Plus 2025 R&P Partisol Model 2300 

http://www..gc.ca/rnspa-naps/Default.asp?lang=En&n=5C0D33CF-1

CAPMoN—Canadian Air and Precipitation Monitoring Network

Canada 2002–Present ECa 29 Rural R&P Partisol Model 2300 PM2.5

Speciation Sampler

http://www.msc.ec.gc.ca/capmon/particulate_general_e.cfm

European Monitoring and Evaluation Program (EMEP)

Norwegian Institute for Air Research

2002–2003 ECa 2 Urban

12 Rural

Thermal Optical Tranmittance – Sunset Lab

http://www.atmos-chem-phys.net/7/5711/2007/acp-7-5711-2007.pdf

European Supersites for Atmospheric Aerosol Research (EUSAAR)

European Union

2006–Present BCa / ECa 20 Rural Aerosol properties including – 

absorption, scattering, AODAethalometer / Sunset Lab

http://www.eusaar.net/files/overview/infrastructures.cfm

China Atmosphere Watch Network (CAWNET)

Chinese Meteorological Administration

1999–Present ECa 6 Urban

12 Rural 

Thermal Optical Reflectance http://www.agu.org/journals/jd/jd08

14/2007JD009525/2007JD009525.pdf
Ambient and Emissions Measurement of Black Carbon

Report to Congress on Black Carbon 253

Worldwide Air Monitoring Networksa, b for Black Carbon

Network Country/ Agency Years of Data BC - Indicator Number of Sites Measurement Method Location of Information and/ or Data

Multiple Independent Sites –two groups by pollutant (BC & ECa) by Vignati et al. (2010)

Multiple Agencies

Various periods 1976–2002

BCaECa

11 Rural7 Rural

Various http://www.atmos-chem-phys.

net/10/2595/2010/acp-10-2595-2010.pdf

a 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.

b 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.

c 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.

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)
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Report to Congress on Black Carbon

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 A1-2.

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. 
Ambient and Emissions Measurement of Black Carbon

Report to Congress on Black Carbon 255

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 A1-3). 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 A1–2 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. 

Table A1-2. Examples of Commercially Available Optical BCa Measurement Techniques. (Source: U.S. EPA)

Instrument (Manufacturer)a Real-time (R) or 

Off-line (O)

Filter (F) or Air 

Stream (A) Wavelengths Measured (See Chapter 2)b

Aethalometer (Magee Scientific) R F 370 nm, 880 nm standard

370, 470, 520, 590, 660, 880 and 950 optional

Particle Soot Absorption Photometer 

(Radiance Research)

R F 467, 530 and 660 nm

Multi-Angle Absorption Photometer 

(Thermo Scientific)

R F 670 nm

Transmissometer (Magee Scientific) O F 370 nm and 880 nm

Densitometer (Tobias Associates Inc.) O F 400 – 650 nm; peak at 575 nm

Smoke Stain Reflectometer (Diffusion 

Systems, Ltd.)

O F Monochromatic light; wavelength not specified

Hybrid Integrating Plate/Sphere O F 633 nm

Photoacoustic soot spectrometer 

(Droplet Measurement Technologies, 

Desert Research Institute)

R A 405, 532, 781 nm

Single particle soot photometer 

(Artium Technologies, Droplet 

Measurement Technologies)

R A 1064 nm

Semi-continuous Field Analyzer (Sunset 

Labs)

R F 632 nm

Photoacoustic Micro Soot Sensor (AVL) R A 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 A1-3 for the MACs used in the comparison studies referenced. 
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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 A1-3 (a) BCa–BCa comparison.

Citation Instrument A Instrument B r r

2 Ratio 

(high/low) Notes

Chow et al. (2009) 7-AE (660 nm) PSAP (660 nm) 0.98 1.28 Fresno Supersite, CA 

Chow et al. (2009) 7-AE (660 nm) MAAP (670 nm) 0.99 3.52 Fresno Supersite, CA

Chow et al. (2009) PSAP (660 nm) MAAP (670 nm) 0.99 2.68 Fresno Supersite, CA

Chow et al. (2009) 7-AE (520 nm) PA (532 nm) 0.96 4.68 Fresno Supersite, CA

Chow et al. (2009) PSAP (530 nm) PA (532 nm) 0.95 3.69 Fresno Supersite, CA

Chow et al. (2009) MAAP (670 nm) PA (670 nm) 0.98 1.51 Fresno Supersite, CA

Snyder and Schauer 

(2007) Aethalometer PSAP 0.93 0.86 1.41 Slope of line (intercept small)

2.68 Overall Average Ratio

Table A1-3 (b) BCa–BCa comparison for study data with correction algorithms applied.

Citation Instrument A Instrument B r r

2 Ratio 

(high/low) Notes

Chow et al. (2009) 7-AE adj (660 nm) PSAP adj (660 nm) 0.95 1.02 Fresno Supersite, CA

Chow et al. (2009) 7-AE adj (660 nm) MAAP (670 nm) 0.97 0.9 Fresno Supersite, CA

Chow et al. (2009) PSAP adj (660 nm) MAAP (670 nm) 0.97 0.81 Fresno Supersite, CA

Chow et al. (2009) 7-AE adj (660 nm) PA (532 nm) 0.95 1.24 Fresno Supersite, CA

Chow et al. (2009) PSAP adj (530 nm) PA (532 nm) 0.95 1.17 Fresno Supersite, CA

1.03 Overall Average Ratio

Table A1-3 (c) ECa–ECa comparisons.

Citation High 

Measurement

Low 

Measurement r n

Ratio 

(High/Low) Notes

Bae et al. (2007) NIER-EC NIOSH TOT UT-EC NIOSH TOT 0.99 223 1.11

Semicontinuous Sunset with 

different temperature protocols: 

NIER - shortened protocols, UT: 

nine-step

Bae et al. (2009) IMPROVE TOR ACE-Asia TOT 0.79 709 2.14 St. Louis Supersite, MO

Cheng et al. 

(2010) IMPROVE_A TOR IMPROVE_A TOT 0.95 89 1.74

TOT and TOR from single DRI Model 

2001 analyzer with denuder

Cheng et al. 

(2010) IMPROVE_A TOR IMPROVE_A TOT 0.95 89 1.83

TOT and TOR from single DRI Model 

2001 analyzer without denuder

Cheng et al. (2011) IMPROVE TOR NIOSH TOT 0.97 81 1.72 Birmingham, AL urban SEARCH site

Cheng et al. (2011) IMPROVE TOR NIOSH TOT 0.92 97 2.78 Jefferson Street, Atlanta, GA urban 

SEARCH site

Cheng et al. (2011) IMPROVE TOR NIOSH TOT 0.85 75 3.57 Pensacola, FL coastal SEARCH site

Cheng et al. (2011) IMPROVE TOR NIOSH TOT 0.81 80 4.00 Centerville, AL rural SEARCH site

Chow et al. (2006) IMPROVE_A TOR STN NIOSH TOT 17 1.87 Fresno Hi-Vol Summer
Ambient and Emissions Measurement of Black Carbon

Report to Congress on Black Carbon 257

Citation High 

Measurement

Low 

Measurement r n

Ratio 

(High/Low) Notes

Chow et al. (2006) IMPROVE_A TOR STN NIOSH TOT 8 2.86 Fresno Hi-Vol Winter

Chow et al. (2006) IMPROVE_A TOR STN NIOSH TOT 17 1.49 Fresno RAAS Summer

Chow et al. (2006) IMPROVE_A TOR STN NIOSH TOT 5 2.17 Fresno RAAS Winter

Chow et al. (2006) IMPROVE_A TOR STN NIOSH TOT 17 1.46 Fresno RAAS Summer with denuder

Chow et al. (2006) IMPROVE_A TOR STN NIOSH TOT 5 1.69 Fresno RAAS Winter with denuder

Chow et al. (2009) IMPROVE_A_TOR_EC IMPROVE_A_TOT EC 0.95 49 1.30 Fresno Supersite, CA

Chow et al. (2009) STN_TOR EC STN_TOT EC 0.9 18 1.41 Fresno Supersite, CA

Chow et al. (2009) IMPROVE_A TOR EC STNTOR EC 0.94 18 1.10 Fresno Supersite, CA

Chow et al. (2009) IMPROVE_A TOR EC French two step EC 0.9 8 1.03 Fresno Supersite, CA

Chow et al. (2009) IMPROVE_A TOR EC Sunset Field EC TOT 0.87 48 1.82 Fresno Supersite, CA

Fujita et al. (2007) IMPROVE TOR NIOSH TOT 0.94 14 1.60 Ambient (urban)

Fujita et al. (2007) IMPROVE TOR NIOSH TOT 0.99 21 1.20 Ambient (on-road)

Gan et al. (2010) IMPROVE TOR NIOSH TOT 6 1.09 Submarine diesel PM

Klouda et al. 

(2005) IMPROVE TOR STN-NIOSH TOT 99 1.66 RM 8785 suspended PM

1.89 Overall Average Ratio

Table A1-3 (d) is on the following two pages.
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Table A1-3 (d) BCa–ECa comparisons.

Citation BCa Method MAC (m2 g-1) l, nm EC Method r Avg BCa Avg EC Ratio BCa/EC Location

Chow et al. (2009) Aethalometer AE-31 PM2.5 16.6 660 IMPROVE_A_TOR PM2.5 0.89 0.94 1.01 0.93 Fresno, CA

Chow et al. (2009) MAAP PM2.5 6.6 670 IMPROVE_A_TOR PM2.5 0.96 0.95 0.95 1.00 Fresno, CAChow et al. (2009) Sunset Optical BC PM2.5 660 IMPROVE_A_TOR PM2.5 0.87 0.52 1.01 0.51 Fresno, CABae et al. (2007) Aethalometer AE-16 PM2.5 16.6 880 Sunset PM2.5 hourly 0.93 0.59 0.68 0.87 Gosan, KoreaBae et al. (2007) Aethalometer AE-16 PM2.5 16.6 880 Sunset PM2.5 hourly 0.92 0.59 0.74 0.80 Gosan, KoreaBae et al. (2007) Aethalometer AE-16 PM2.5 16.6 880 Sunset PM2.5 hourly 0.80 1.89 2.18 0.87 Gosan, KoreaBae et al. (2007) Aethalometer AE-16 PM2.5 16.6 880 Sunset PM2.5 hourly 0.70 1.89 2.3 0.82 Gosan, KoreaJeong et al. (2004) Aethalometer AE-20 16.6 880 Sunset PM2.5 every two hrs. 0.92 0.9a 0.4a 2.25 Rochester, NY

Jeong et al. (2004) Aethalometer AE-20 16.6 880 Sunset PM2.5 every two hrs. 0.77 0.9 0.4a 2.25 Philadelphia, PAJeong et al. (2004) Sunset Optical BC PM2.5 16.6 660 Sunset PM2.5 every two hrs. 0.97 0.3a 0.4a 0.75 Rochester, NYJeong et al. (2004) Sunset Optical BC PM2.5 16.6 660 Sunset PM2.5 every two hrs. 0.85 0.4a 0.4a 1.00 Philadelphia, PA

Hagler et al. (2007) PSAP b 565 NIOSH TOT 0.95 7ngm-3 b Greenland - no BC mass

Hitzenberger (2006) AE-9 19 830 Cachier two step 1000C in O2 0.72 1.14 Vienna, AustriaHitzenberger (2006) MAAP 6.5 670 Cachier two step 1000C in O2 0.91 1.20 Vienna, Austria

Hitzenberger (2006) Integrating sphere Calcc 550 Cachier two step 1000C in O2 0.86 0.98 Vienna, AustriaHitzenberger (2006) Light transmission-white light white Cachier two step 1000C in O2 0.89 1.20 Vienna, Austria

Hitzenberger (2006) AE-9 19 830 VDI 650 C in O2 0.66 0.95 Vienna, Austria

Hitzenberger (2006) MAAP 6.5 670 VDI 650 C in O2 0.88 1.05 Vienna, AustriaHitzenberger (2006) Integrating sphere Calcc 550 VDI 650 C in O2 0.78 0.84 Vienna, Austria

Hitzenberger (2006) Light transmission-white light white VDI 650 C in O2 0.79 1.05 Vienna, Austria

Hitzenberger (2006) Aethalometer AE-9 19 830 TOT 800C in O2 0.61 1.11 Vienna, AustriaHitzenberger (2006) MAAP 6.5 670 TOT 800C in O2 0.88 1.11 Vienna, Austria

Hitzenberger (2006) Integrating sphere Calcc 550 TOT 800C in O2 0.67 0.93 Vienna, AustriaHitzenberger (2006) Light transmission-white light white TOT 800C in O2 0.83 1.13 Vienna, Austria

Snyder and Schauer (2007) PSAP b 565 Sunset PM2.5 hourly 0.91 b Riverside, CA 

Snyder and Schauer (2007) Aethalometer AE-31 b 880 Sunset PM2.5 hourly 0.93 b Riverside, CA Sharma et al. (2002) Aethalometer AE-11 PM2.5 19 880 IMPROVE TOR PM2.5 0.89 0.58a Egbert, Canada

Sharma et al. (2002) PSAP PM2.5 10 565 IMPROVE TOR PM2.5 0.99 0.58a Egbert, CanadaSharma et al. (2002) Aethalometer AE-11 PM2.5 19 880 IMPROVE TOR PM2.5 0.98 1.42a Downsview, Canada
Ambient and Emissions Measurement of Black Carbon

Report to Congress on Black Carbon 259

Citation BCa Method MAC (m2 g-1) l, nm EC Method r Avg BCa Avg EC Ratio BCa/EC Location

Sharma et al. (2002) PSAP PM2.5 10 565 IMPROVE TOR PM2.5 0.69 1.42a Downsview, CanadaSharma et al. (2002) Aethalometer AE-11 19 880 Cachier two step EC 1100C in O2 0.91 0.087, 0.012a Alert, Canada (Arctic)Sharma et al. (2002) PSAP 10 565 Cachier two step EC 1100C in O2 0.93 0.087, 0.012a Alert, Canada (Arctic)

Sharma et al. (2002) Aethalometer AE-11 PM2.5 19 880 NIOSH TOT PM2.5 0.92 1.95a Evans Ave, CanadaSharma et al. (2002) PSAP PM2.5 10 565 NIOSH TOT PM2.5 0.96 1.95a Evans Ave, Canada

Sharma et al. (2002) Aethalometer AE-11 PM2.5 19 880 NIOSH TOT PM2.5 0.89 1.82a Palmerston, Canada

Sharma et al. (2002) PSAP PM2.5 10 565 NIOSH TOT PM2.5 0.89 1.82a Palmerston, CanadaSharma et al. (2002) Aethalometer AE-11 PM2.5 19 880 NIOSH TOT PM2.5 0.92 1.48a Winchester, Canada

Sharma et al. (2002) PSAP PM2.5 10 565 NIOSH TOT PM2.5 0.54 1.48a Winchester, CanadaVenkatachari et al. (2006) Aethalometer AE-20 PM2.5 16.6 880 Sunset PM2.5 hourly n/a 1.01 0.85 1.2 New York City, NYVenkatachari et al. (2006) Aethalometer AE-20 PM2.5 16.6 880 R&P 5400 PM2.5 hourly n/a 1.01 0.55 1.8 New York City, NYVenkatachari et al. (2006) Aethalometer AE-20 PM2.5 16.6 880 CSN TOT PM2.5 n/a 1.01 0.53 1.9 New York City, NYSahu et al. (2009) PSAP PM2.5 8.9 565 Sunset PM2.5 hourly 0.98 1.18 n/a 1.0 Jeju Island, South KoreaYang et al. (2006) Aethalometer AE-16 PM2.5 16.6 880 IMPROVE TOR PM2.5 0.72 16.5 12 1.4 Xi’an, ChinaMiyazaki et al. (2008) COSMOS 9.8 565 Sunset PM1.0 hourly 0.96 n/a n/a n/a Thailand

Babich et al. (2000) Aethalometer AE-20 19 880 IMPROVE TOR PM2.5 0.87 1.1 1.4 0.79 Bakersfield, CA

Babich et al. (2000) Aethalometer AE-20 19 880 IMPROVE TOR PM2.5 0.98 1.2 1.5 0.80 Chicago, ILBabich et al. (2000) Aethalometer AE-20 19 880 IMPROVE TOR PM2.5 0.95 0.8 1.3 0.62 Dallas, TXBabich et al. (2000) Aethalometer AE-20 19 880 IMPROVE TOR PM2.5 0.95 1.1 1.5 0.73 Philadelphia, PA

Babich et al. (2000) Aethalometer AE-20 19 880 IMPROVE TOR PM2.5 0.96 3.1 3.9 0.79 Phoenix, AZBabich et al. (2000) Aethalometer AE-20 19 880 IMPROVE TOR PM2.5 0.92 1.6 1.9 0.84 Riverside,CARam et al. (2010) Aethalometer 16.6 880 Sunset TOT NIOSH PM10 0.79 4.45 3.84 1.2 Kanpur, IndiaHusain et al. (2007) Aethalometer AE-21 PM3.2 16.6 880 Sunset TOT NIOSH PM2.5 0.84 n/a n/a 1.3 Lohore, PakistanLim et al. (2003) PSAP PM2.5 10 565 R&P 5400 PM2.5 hourly n/a 1.26 2.8 0.5 Atlanta, GALim et al. (2003) PSAP PM2.5 10 565 RU/OGI TOT PM2.5 hourly Sunset predecessor n/a 1.26 2.33 0.5 Atlanta, GA

Lim et al. (2003) Aethalometer AE-16 PM2.5 12.6 880 R&P 5400 PM2.5 hourly n/a 2.61 2.8 0.9 Atlanta, GALim et al. (2003) Aethalometer AE-16 PM2.5 12.6 880 RU/OGI TOT PM2.5 hourly Sunset predecessor n/a 2.61 2.33 1.1 Atlanta, GA

a Median concentration.b BC data presented as absorption coefficients (Mm-1). 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.
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Instruments measuring light absorption are often 

capable of measuring light absorption at additional 

wavelengths in the near-UV or UV, which may help 

indicate the presence of BrC. Such approaches are 

currently used to attempt separation of the influence 

of wood smoke aerosols, which tend to be rich 

in BrC, from those that are dominated by diesel 

emissions and other fossil fuel combustion, which are 

rich in BC.

The disagreement among BCa measurements may 

be due in part to differing instrument sensitivities 

and responses to other PM components (Slowik et 

al., 2007), filter-loading artifacts, or the use of an 

incorrect light-absorption-to-BC mass concentration 

conversion factor for studies reporting BCa in terms 

of their mass concentrations. Chow et al. (2009) 

found that applying post-processing algorithms 

greatly improved the agreement among different 

filter-based BCa techniques (Refer to Table A1-3 

above). EPA and other researchers (Collaud Coen et 

al., 2010; Park et al., 2010; Virkkula et al., 2007; Bond 

et al., 1999; Chen et al., 2004; Jimenez et al., 2007) 

are similarly assessing whether post-processing 

algorithms and site-specific conversion factors 

may also be beneficial to better understand the 

differences among BCa and ECa measurements. 

In addition to issues with the various BCa and ECa 

measurement techniques mentioned above, BCa-toECa ratios are likely to be affected by the presence 

of other light-absorbing species (e.g., BrC and 

dust). The specific inter-comparison circumstances 

(e.g., location, season, sample collection and 

analysis procedures, optical wavelength, data 

corrections, and aerosol mixture1F ) may be 

important to understand and reconcile reported 

differences. A summary of the data presented in 

Figure A1-2 and comparisons of BCa/BCa (with 

and without correction algorithms) and ECa/

ECa along with the circumstances for the intercomparison measurements is found in Table A1-3 

(Inter-comparison of Ambient BCa and ECa 

Measurements). 

A1.2.4 Inter-comparisons Among ThermalOptical ECa Measurements 

The differentiation between OCa and ECa by 

thermal-optical methods is highly dependent on 

the specific measurement method operational 

procedures used. Different instrument operating 

procedures, thermal temperature profiles, and 

correction for char can lead to differences in ECa. 

Previous method comparison studies summarized 

1

 The mixture can be important because of the relative amount of 

non-graphitic absorbing materials, e.g., BrC and dust, as well as 

internal mixtures with water, organics, and sulfates.

Figure A1-2. Reported BCa/ECa Ratios for a Pair of Measurement Techniques Reported in Ambient 

Field Studies. (Source: U.S. EPA)
Ambient and Emissions Measurement of Black Carbon

Report to Congress on Black Carbon 261

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 A1-3 (Klouda et al., 2005; Bae et 

al., 2007; 2009; Cheng et al., 2010; 2011; Chow et al., 

2006; 2009; Fujita et al., 2007; Gan 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 A1-3). 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 NIOSHlike TOT method in Figure A1-4 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 
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in Figure A1‐3. 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). 

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 A1-5 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 μg/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 

2

 11 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.

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)
Ambient and Emissions Measurement of Black Carbon

Report to Congress on Black Carbon 263

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 μg/

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)

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 AprilSeptember are shown in red. The dotted line is a spline fit through the monthly data. (Source: U.S. EPA, AQS)
264

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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 

A1-6 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 A1-7.) The reported EC 

data based on the NIOSH-type 

measurements in Figure A1-6 

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 A1‐4, 

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 

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_A TOR monitoring protocol is 

shown in blue. (Source: U.S. EPA, AQS) 

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)
Ambient and Emissions Measurement of Black Carbon

Report to Congress on Black Carbon 265

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 filterloading-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:

y 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.

y Having a consistent, well-defined “BC” reference 

material would help to better understand method 

differences and define the uncertainties in the 

various measurement methods.

y 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.

y 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 nearUV and UV wavelengths.

y To ensure proper use of measurements, 

consistent data reporting (including metadata) 

of the sampling and analysis protocols and data 

adjustments must be provided. 

y Continued research and further development 

is needed for continuous or real-time single 

particle measurements (e.g., aerosol time-offlight 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 
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Report to Congress on Black Carbon

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 PM10 size fractions. 

Some local/state and site specific standards also 

require testing for PM10 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 PM Type Filtration 

Temperature (°F) Purpose CFR Reference

EPA Method 5 Filterable 248 ± 25 General 40 CFR 60 Appendix A-3

EPA Method 5A Filterable 108 ± 18 Asphalt Roofing 40 CFR 60 Appendix A-3

EPA Method 5B Filterable 320 ± 25 Utility Plants 40 CFR 60 Appendix A-3

EPA Method 5D Filterable 248 ± 25 Positive Pressure baghouses 40 CFR 60 Appendix A-3

EPA Method 5E Filterable and Total 

Organic Material 248 ± 25 Wool Fiberglass 40 CFR 60 Appendix A-3

EPA Method 5F Filterable 320 ± 25 Non sulfate Filterable PM 40 CFR 60 Appendix A-3

EPA Method 5G Filterable and 

Condensable <90 Wood Heaters - Dilution 40 CFR 60 Appendix A-3

EPA Method 5H Filterable and 

Condensable <248 and <68 Wood Heaters 40 CFR 60 Appendix A-3

EPA Method 5I Filterable 248 ± 25 Low level general 40 CFR 60 Appendix A-3

EPA Method 17 Filterable Stack Temperature General 40 CFR 60 Appendix A-6

EPA Method 201 Filterable

10 μm Stack Temperature General – Particle Sizing 40 CFR 51 Appendix M

EPA Method 201A Filterable

10 μm/2.5 μm Stack Temperature General – Particle Sizing 40 CFR 51 Appendix M

EPA Method 202 Condensable 85 General – Condensable PM 40 CFR 51 Appendix M

EPA Conditional Test 

Method -039

Total 10 μm/2.5 μm 

(Filterable

and Condensable)

85 General – Dilution based PM

Example State, VCS, and International Methods

CARB 5 Filterable 248 ± 25

CARB 501 Filterable, multiple 

aerodynamic sizes Stack Temperature General – Particle Size http://www.arb.ca.gov/testmeth/vol1/Meth_501.pdf

ASTM D6831 - 05a Filterable Stack Temperature Continuous PM

ISO 9096 and EN 13284 Filterable

VDI 2066 Part. 10 

method and in the Norm 

EN 13284-1

Filterable 

10 μm/2.5 μm
Ambient and Emissions Measurement of Black Carbon

Report to Congress on Black Carbon 267

measurement protocol (U.S. EPA, 2010f); over time 

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 PM10 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 A1-4 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 lbs. 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 at 45%RH, while ambient 

measurements and many other source tests use 35%RH. 
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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 A1-5, 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 PM2.5

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 PM2.5 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 PM2.5 estimation methods based on PM10

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 PM10 to generate estimates of PM2.5

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 PM2.5 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.
Ambient and Emissions Measurement of Black Carbon

Report to Congress on Black Carbon 269

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

y Understand how the source EC values relate to 

source BC values based on currently available 

techniques.

y 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.

y Develop a standard “BC” reference material and 

establish a standard measurement method to 

report source data as BC.

2. Stationary source measurements

y 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.

y 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.

y 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

y Develop standard measurement methods for BC 

for both on-road and nonroad engines, especially 

diesels but also gasoline vehicles/engines. 

y 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.
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Appendix 1

Report to Congress on Black Carbon Stats (%ile) Agricultural Burning Bituminous Combustion Charbroiling Distillate Oil Combustion HDDV Exhaust LDDV Exhaust Natural Gas Combustion Noncatalyst Gasoline Exhaust Onroad Gasoline Exhaust Prescribed Burning ProcessGas Combustion Residential Wood Combustion: HardSoft SubBituminous Combustion Wildfires Wood Fired Boiler OC Fraction of PM2.5 10th 0.30 0.02 0.34 0.25 0.18 0.20 0.25 0.84 0.30 0.65 0.05 0.39 0.02 0.47 0.33 25th 0.34 0.02 0.46 0.25 0.18 0.25 0.25 0.84 0.44 0.70 0.20 0.47 0.02 0.47 0.33 50th 0.40 0.03 0.70 0.25 0.18 0.32 0.25 0.84 0.55 0.71 0.35 0.53 0.03 0.56 0.33 75th 0.44 0.07 0.84 0.25 0.18 0.39 0.25 0.84 0.67 0.79 0.46 0.58 0.04 0.64 0.33 90th 0.56 0.12 0.87 0.25 0.18 0.44 0.25 0.84 0.75 0.83 0.57 0.68 0.04 0.64 0.33 N 9 3 4 1 1 4 1 1 10 7 3 12 2 2 1 BC Fraction of PM2.5 10th 0.05 0.01 0.00 0.10 0.77 0.31 0.38 0.01 0.09 0.01 0.10 0.01 0.02 0.03 0.14 25th 0.08 0.01 0.01 0.10 0.77 0.38 0.38 0.01 0.14 0.01 0.13 0.04 0.02 0.03 0.14

50th 0.10 0.02 0.02 0.10 0.77 0.53 0.38 0.01 0.19 0.02 0.17 0.06 0.04 0.09 0.1475th 0.12 0.05 0.06 0.10 0.77 0.63 0.38 0.01 0.23 0.04 0.22 0.10 0.07 0.16 0.1490th 0.13 0.08 0.10 0.10 0.77 0.64 0.38 0.01 0.34 0.07 0.27 0.12 0.07 0.16 0.14N 9 3 3 1 1 4 1 1 10 7 3 12 2 2 1OC:BC Ratios10th 5.9 1.9 2.5 0.2 0.6 0.6 59.9 3.3 54.3 0.5 27.6 1.0 14.5 2.425th 4.2 1.7 41.2 2.5 0.2 0.7 0.6 59.9 3.2 49.4 1.5 12.4 1.0 14.5 2.450th 4.1 1.6 31.1 2.5 0.2 0.6 0.6 59.9 2.9 38.6 2.1 9.4 0.7 5.9 2.475th 3.6 1.5 13.5 2.5 0.2 0.6 0.6 59.9 3.0 19.3 2.1 5.9 0.7 4.1 2.490th 4.3 1.4 8.5 2.5 0.2 0.7 0.6 59.9 2.2 12.0 2.1 5.5 0.7 4.1 2.4BC+OC, as % PM10th 36% 3% 34% 35% 95% 51% 63% 85% 39% 66% 15% 40% 4% 50% 46%25th 43% 4% 47% 35% 95% 63% 63% 85% 57% 72% 34% 51% 4% 50% 46%50th 49% 4% 72% 35% 95% 86% 63% 85% 74% 73% 53% 58% 7% 65% 46%75th 57% 12% 90% 35% 95% 102% 63% 85% 90% 83% 68% 68% 11% 80% 46%

90th 69% 20% 97% 35% 95% 108% 63% 85% 109% 90% 83% 81% 11% 80% 46%

Table A1-5. 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)
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 PM2.5. PM2.5 emission factors 

and the speciation factors for particular constituents 

Appendix 2

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 PM2.5 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
272

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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 = A x EF (1-ER/100) (Equation 1)

Where

y E = PM2.5 emissions (for example, in Tons);

y A = activity rate;

y EF = the emission factor, and

y 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 PM2.5 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 

1 http://cfpub.epa.gov/webfire.

set of conditions. Under some circumstances in 

the inventorying of PM2.5, 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 PM2.5 for many stationary 

and area source categories. This introduces some 

uncertainties in the emissions estimates.
Black Carbon Emissions Inventory Methods and Comparisons

Report to Congress on Black Carbon 273

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 
274

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Report to Congress on Black Carbon

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:

y 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). 

y 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. 

y 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).

y Finally, many “augmentations” are done in the 

emissions inventory processing steps. These 

augmentations include scaling measured PM to 

PM2.5 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:

y On-road gasoline, such as passenger cars and 

light-duty trucks 

y 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.

y Nonroad diesel, including construction, 

agricultural, and other equipment

y Nonroad gasoline, including both 2-stroke and 

4-stroke cycle engines such as those used in 

lawn/garden equipment and recreational marine

y Commercial marine, classified by engine 

displacement as categories C1, C2, and C3 

(ocean-going)

y Locomotives

y 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/models/moves/420r10026.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/models/moves/movesback.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/models/moves/movesback.htm.
Black Carbon Emissions Inventory Methods and Comparisons

Report to Congress on Black Carbon 275

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 EPA 420-R-05-016 Exhaust Emission 

Effects of Fuel Sulfur and Oxygen on Gasoline Nonroad Engines, 

http://www.epa.gov/oms/nonrdmdl.htm#model.

EPA (July 2010) Report EPA-420R-10-017 Nonroad Engine Population 

Estimates, http://www.epa.gov/oms/nonrdmdl.htm#model.

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/nonrdmdl.htm#model.

EPA (July 2010) EP{A Report 420-R-10-015 Median Life, Annual 

Activity, and Load Factor Values for Nonroad Engine Emissions 

Modeling, http://www.epa.gov/oms/nonrdmdl.htm#model.

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/nonrdmdl.htm#model.

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.htm#model.

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.htm#model.

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 
276

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Report to Congress on Black Carbon

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 

EPA’s 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 noncatalyst 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 C1/

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 
Black Carbon Emissions Inventory Methods and Comparisons

Report to Congress on Black Carbon 277

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 (thermaloptical 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 populationweighted 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: slowspeed diesel (SSD) and medium speed diesel (MSD). 

Lack et al. (2009) also grouped the BC emission 

observations according to fuel sulfur content. Fiftyone ship observations had low sulfur fuel content 

(<5,000 ppm) and 42 ship observations of vessels 

Table A2-1. Summary of Recent Studies that Measured BC and PM Emission Rates from C3 Marine. 

(Source: U.S. EPA)

Study Vessel Fuel Fuel Sulfur 

Content BC/PMa

Murphy et al. (2009) Post-Panamax Container Heavy Fuel Oil 30,000 ppm 0.31%

Agrawal et al. (2008) Suezmax Marine Tanker Heavy Fuel Oil 28,500 ppm 0.50%

Petzold et al. (2010)b Medium Speed Diesel Engine Heavy Fuel Oil 22,100 ppm 2.63%

Lack et al. (2009) Slow-speed diesel vesselsc Variety Varietyd 7.33%

Lack et al. (2009) Medium-speed diesel vesselse Variety Varietyf 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.

d 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).

Reference BC EC OC or OM BC Measurement EC/OC Measurement

Agrawal et al. (2008) 0.5% 11% Thermal optical transmittance 

(TOT) NIOSH 5040 Murphy et al. (2008) 0.3% 7.4%

Petzhold et al. (2010)

0.85% 2.6% 21%

Multi-Angle Absorption 

Photometer, MAAP

Thermal-Optical Method,VDI 2465 

Part 2

(SSD) Lack et al. (2009) 7.3% 23% Photoacoustic Aerosol 

Spectrometer with 

adsorption at 532 nm

Organic Matter (OM) calculated by 

subtracting SO4,NO3, and NH4 from 

aerosol mass spectrometer (AMS) 

measurements

(MSD) Lack et al. (2009) 28% 16%

(>5,000 ppm) Lack et al. (2009) 7.1% 19%

(<5,000 ppm) Lack et al. (2009) 51% 35%

Table A2-2. Summary of Speciation Ratios of PM from Relevant Marine Studies. (Source: U.S. EPA)
278

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Report to Congress on Black Carbon

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 (EI) 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 nonLTO 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 EI for PM BC 

6

 The FAA SAGE website is http://www.faa.gov/about/office_org/

headquarters_offices/apl/research/models/sage.
Black Carbon Emissions Inventory Methods and Comparisons

Report to Congress on Black Carbon 279

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 insitu 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 EI 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 EIs 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 

Table A2-3. Estimates of Aircraft BC Emissions. (Source: U.S. EPA)

Reference Aircraft Engine(s) Measurement Condition BC EI (g/kg fuel)

Kinsey et al. (2011) DC8

Various air 

frames APEX 1 

to 3

CFM56-2C1

CFM56-7B24

CFM56-3B1/3B2

CFM56-3B1

RB211-535E4B

Non-LTO thrust levels at Sea Level 

Static (SLS)

0.021, 0.026, 0.032

0.028, 0.025

0.092

0.098

0.275

Petzold et al. (1999a) ATTAS Rolls-Royce/SNECMA 

M45HMk501

Non-LTO thrust levels at SLS

In-flight

0.118-0.149

0.11-0.15

Pueschel et al. (1997) Concorde Olympus 593 In-flight 16300 m altitude 0.07-0.11

Petzold et al. (1999b) B737-300 CFM56-3B1 In-flight 7925 m altitude 0.01

Petzold et al. (1999b) A310-300 CF6-80C2A2 In-flight 10670 m altitude 0.021

Petzold et al. (1999b) VFW 614 M45H In-flight 7925 m altitude 0.07-0.11

Lilenfield et al. (1995) DC8 GE 404 All thrust levels at SLS 0.03-0.4

Anderson et al. (1998) Multiple Multiple In-flight; mass EIs estimated from 

number EIs based on average 

particle volume and mass 

0.01-0.35

Dopelheuer (2001) n/a CF6-50C2 Modeled Cruise Simulation - DLR 

Method using empirical calculation

0.015
280

Appendix 2

Report to Congress on Black Carbon

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 et al. (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. Countryspecific fuel consumption rates were compiled 

for industrial, domestic, and combined sectors 

for solid, liquid, and gaseous fuels. Country-wide 

Figure A2-1. Bond et al. Methodology for Developing 

Emissions Estimates. (Source: Bond et al., 2004)
Black Carbon Emissions Inventory Methods and Comparisons

Report to Congress on Black Carbon 281

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:

y 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.

y 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 
282

Appendix 2

Report to Congress on Black Carbon

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 

Informationa

Reddy and 

Venkataraman 

(2002b, Fossil 

Fuel) Reddy and 

Venkataraman 

(2002a, Biomass 

Combustion) Black 

Carbon and Organic 

Carbon

India, 1996-1997 

(.25 x .25) 

4 utilities

5 coal combustion

8 industrial

2 residential/ commercial

8 transportation

4 biomass/biofuels burning

Literature review 

U.S. EPA AP-42 

Customized 

emission factors 

to fit Indian 

technology

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.

Streets et al. (2003) 

Black Carbon Only

Asia, 2000 

(1 x 1 to .08 x .08)

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.

Literature review 

U.S. EPA AP-42 

MOBILE5 model

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.

Cao et al. (2006) Black 

Carbon and Organic 

Carbon

China, 2000 

(.2 x .2)

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.

Literature review 

Laboratory tests of 

biofuel emissions 

from cooking 

stoves

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.

Streets et al. (2001) 

Black Carbon 

China, 1995 Covers 37 different source 

types over five sectors 

(power, industry, residential, 

transport, field combustion) 

and 13 different fuel types, 

including biofuels

Literature Review Fuel consumption data by sector and 

fuel type were developed within the 

framework of the RAINS-Asia model 

(Hordijk et al., 1995). Generated by 

China’s Energy Research Institute.

Parashar et al. (2005) 

Black Carbon and 

Organic Carbon

India, 1995 Fossil fuel, biofuel, and 

biomass burning

Literature Review

Laboratory tests of 

biofuels and soft 

coke

Fossil fuel consumption from TEDDY 

2001/2002, biofuel use and biomass 

burning taken from Reddy and 

Venkataraman (2002b).

Sahu et al. (2008) 

Black Carbon

India, 2001

(1 x 1)

Categorized into area sources 

and LPS and then by fuel 

type

Literature review Activity data collected from Central 

Electricity Authority (CEA), Census 

of India, Ministry of Coal, Ministry 

of Road Transport and Highways, 

Ministry of Agriculture.

Dickerson et al. 

(2002) Black Carbon

Asia, 1991 and 

2001

Same as 37 different source 

types identified in Streets et 

al. (2001)

Literature Review

MOBILE 5, also 

used CO/BC 

ratio to estimate 

emissions

RAINS-Asia model, Tata Energy 

Research Institute.
Black Carbon Emissions Inventory Methods and Comparisons

Report to Congress on Black Carbon 283

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).

y 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. 

y 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.

y 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 

Carbon Emission 

Inventory 

Reference

Region, 

Base Year 

(Resolution, if 

available)

Emission Source 

Categories

Source of 

Emission Factors

Source(s) of Activity 

Informationa

Lamarque et al. 

(2010) Black Carbon 

and Organic Carbon

Global, 2000

(.5 x .5)

12 different sectors over 40 

different regions 

Literature Review 

(Junker and 

Liousse, 2008 

combination 

emission factors) 

Biomass from RETRO, GICC, and 

GFEDv2 inventories; ship data from 

International Maritime Organization 

(IMO); aircraft data from AERO2K 

database, EUROCONTROL, Bond et 

al. (2007), and Junker and Liousse 

(2008).

Mitra et al. (2002) 

Black Carbon and 

Organic Carbon

India, 1996 7 different fuel types Literature Review 

(from Cooke and 

Wilson, 1996)

Mayol-Bracero et al. 

(2002) Black Carbon

India, 2000 8 different fuel types across 

four (five?) sectors

Literature Review 

(Streets et al., 

2001) )

RAINS-Asia model.

Ohara et al. (2007) 

Black Carbon and 

Organic Carbon

Asia, 2000 (.5 x .5) Four sectors broken down 

into fuel type (coal, oil, 

biofuel, others)

Literature Review 

U.S. EPA AP-42

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.

Derwent et al. (2001) Western 

Europe,1995-1998

All sectors Back calculated 

using dispersion 

modeling and 

ambient data

NA

a Unless otherwise noted, refer to main Emission Inventory Reference for a list of specific publications and databases from which underlying 

data were drawn.
284

Appendix 2

Report to Congress on Black Carbon

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.

y 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 underdeveloped country emission factors for all fossil 

fuel combustion sectors and also from use of 

diesel activity information which did not represent 

current conditions.

y 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.

y 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.

y 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.”

y 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.
Studies Estimating Global and 

Regional Health Benefits of 

Reductions in Black Carbon

Appendix 3

Geographic 

Scale Results of Study Mitigation Measures Pollutants Reference

Studies of Mitigation Strategies for Ambient Reductions in BC

Global, Arctic 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.

Elimination of global 

anthropogenic FS and BSG.

PM2.5 from fossil 

fuel soot, biofuel 

soot, & methane

Jacobson 

(2010)

Global 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.

50% reduction in China’s 2030 

SO2, OC, BC emissions from 2000 

levels.

SO2, OC, BC Saikawa et al. 

(2009) 

Global 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.

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.

SO2, OC, BC, NO3, 

O3

Shindell et al. 

(2011) 

Global 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.

50% reductions in anthropogenic 

BC emissions globally, from 

8 major world regions, and 

from 3 major economic 

sectors (residential, industrial, 

transportation). 

BC, OC Anenberg et al. 

(2011) 

Global Implementing all measures would 

avoid 1-5 million PM2.5 and O3-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 lowcarbon CO2 measures.

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 highemitting vehicles, banning open 

burning of agricultural waste, and 

eliminating biomass cook stoves 

in developing countries).

SO2, OC, BC, NO3, 

O3

UNEP and 

WMO (2011a) 

Report to Congress on Black Carbon 285
286

Appendix 3

Report to Congress on Black Carbon

Geographic 

Scale Results of Study Mitigation Measures Pollutants Reference

Studies of Indoor and Ambient Mitigation Strategies for BC

China, India, 

Africa

Benefits of mitigation exceeded costs by 

factors of 3.6 to 13.6 to one.

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.

BC, OC, SO2 Baron et al. 

(2009)

China, India 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.

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, SO2 Kandlikar et al. 

(2009)

Studies of Indoor Mitigation Strategies to Reduce BC

China Climate and human health benefits to 

cost ratio of 6 with about 69% of these 

benefits associated with human health.

Household fuel intervention. BC, OC, SO2 Smith et al. 

(2008) 

India and UK 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. 

Energy efficiency in UK household 

heating and 150 million lowemission cookstoves in India.

BC, OC, sulfates Wilkinson et al. 

(2009)
Efforts to Limit Diesel Fuel Sulfur 

Levels

As discussed in Chapter 8, the availability of lowsulfur 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 inuse 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:

y Africa: Morocco established limits of 50 ppm in 

2009, and Tunisia will require 50 ppm fuel in 2014-

2015.

y 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).

y 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.

y East Asia and Pacific Islands: Malaysia required 

50 ppm diesel fuel in 2010 and is requiring 

Appendix 4

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.

y Eastern Europe: lbania 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.

y 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.

y 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
288

Appendix 4

Report to Congress on Black Carbon

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.
Efforts to Limit Diesel Fuel Sulfur Levels

Report to Congress on Black Carbon 289

Table A4-1. International Regulations and International Agreements on Diesel Fuel Sulfur Levels (in ppm). (Source: U.S. EPA)

Region Country Year Current Maximum Level 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 2018 2019 2020

Americas and Caribbean

Mexico (National) ? ? ? ? ? ? 500 500 500 15 15Mexico (Northern) ? ? ? ? ? ? 300 15 15 15 15Argentina (Urban) ? ? ? ? ? ? 1500 1500 1500 1500 (500) (500) 50 50

Argentina (NonUrban) ? ? ? ? ? ? 2500 2500 2500 2500 2500 2500 1500 1500Barbados No existing or planned standards. (5000)Bolivia ? ? ? ? ? ? ? ? ? ? (2000) 5000Brazil (NonUrban) ? 3500 3500 3500 3500 3500 3500 3500 3500 3500 1800 1800 1800 1800 500 2000Brazil (Metropolitan) ? 2000 2000 2000 2000 500 500 500 500 50 50 50 50 10 50Chile (National) ? ? ? ? ? ? 350 350 350 350 50 50Chile (Santiago) ? ? ? ? ? ? 50 50 50 50 10 10Colombia (National) ? ? ? ? ? ? 2500 2500 2500 2500 500 50 500Colombia (Bogota) ? ? ? ? ? ? ? ? ? ? 50Costa Rica No existing or planned standards. (500)Cuba No existing or planned standards. (8000)Dominican Republic No existing or planned standards. 7500Ecuador (National) ? ? ? ? ? ? ? (500) 500 500 50Ecuador (Urban) ? ? ? ? ? ? 5000 500Ecuador (NonUrban) ? ? ? ? ? ? 7000El Salvador ? ? ? ? ? ? ? ? ? ? 500Guatemala ? ? ? ? ? ? ? ? ? ? (500) 5000Honduras ? ? ? ? ? ? ? ? ? ? (500) 5000Panama ? ? ? ? ? ? ? ? ? ? (1000) 5000Peru (Urban) ? ? ? ? ? ? 1500 1500 1500 1500 50
290

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Report to Congress on Black Carbon

Region Country Year Current Maximum Level 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 2018 2019 2020

Americas and 

Caribean

Peru (NonUrban) ? ? ? ? ? ? 3000 3000 3000 3000 1500Uruguay ? ? ? ? ? ? 8000 8000 8000 8000 50Venezuela ? ? ? ? ? ? 5000 5000 5000 5000 2000 5000

East Asia, Australia, and Pacific Islands

Australia ? ? ? 500 500 500 50 50 50 10 10Cambodia 2000 ? ? ? ? ? 1500China (National) 2000 2000 2000 2000 2000 2000 2000 2000 2000 2000 350China (Beijing) ? ? ? ? ? ? ? ? 50 50 50 50 10China (Hong Kong) 500 50 50 50 50 50 50 50 50 50 50Fiji ? ? ? ? ? ? ? 500Indonesia 5000 5000 5000 5000 5000 5000 5000 3500 3500 3500 (500) 4000Japan 100 100 100 100 100 50 50 50 50 50 50Malaysia 3000 3000 (500) (500) (500) (500) (500) (500) (500) (500) 50 50 50 50 50 10Nepal 10000 10000 10000 10000 10000 10000 10000 10000 10000 10000 10000 500Philippines ? ? ? ? ? ? ? ? ? ? 500Republic of Korea 500 500 500 500 500 500 500 500 500 500 50Singapore 500 500 500 500 500 500 500 50 50 50 50Thailand 500 500 500 500 500 500 500 500 500 500 500 500 50Vietnam 10000 10000 2000 2000 2000 500 500 500 500 500 500

Eastern Europe, Caucasus, and Central Asia

Albania ? ? ? ? ? ? ? ? ? ? 350 150 10 2000Belarus ? ? ? ? ? ? ? ? ? ? 50 10 350Bosnia & Herzegovina ? ? ? ? ? ? ? ? ? ? 350 350Croatia ? ? ? ? ? ? ? ? ? 50 50 50 10Russia ? ? ? ? ? ? ? ? 500 350 50Turkey ? ? ? ? ? ? ? ? 50/ 1000 50/ 1000 50/ 1000Armenia ? ? ? ? ? ? ? ? 350 50 10Azerbaijan ? ? ? ? ? ? ? ? ? ? 2000 2000 2000 2000 2000 500Georgia ? ? ? ? ? ? ? ? ? 350 50Kazakhstan ? ? ? ? ? ? ? ? 2000 2000 10Kyrgyzstan ? ? ? ? ? ? ? ? (350) (350) (350) 2000- 5000Serbia ? ? ? ? ? ? ? ? ? 10000 (350) 10000Uzbekistan ? ? ? ? ? ? ? ? ? ? 5000 400- 2000
Efforts to Limit Diesel Fuel Sulfur Levels

Report to Congress on Black Carbon 291

Region Country Year Current Maximum Level 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 2018 2019 2020

South Asia

Afghanistan No existing or planned standards. Bangladesh ? ? ? ? ? ? 5000India 2500 2500 2500 2500 2500 500 500 500 500 500 350India (Selected Areas) 350 350 350 350 350 50Nepal 10000 10000 10000 10000 10000 10000 (500) (500) (500) (500) (500) (350)Pakistan ? ? ? ? ? ? ? ? ? ? ? ? 500Sri Lanka 10000 10000 10000 3000 ? ? ? ? ? ? 500

Southwest Asia and North Africa

Algeria No existing or planned standards. 900Bahrain ? ? ? ? ? ? ? ? ? ? (500) 500Egypt ? ? ? ? ? ? ? ? ? ? (5000) 10000Iran ? ? ? ? ? ? ? ? ? ? (5000) 10000Iraq ? ? ? ? ? ? ? ? ? ? (10000) 25000Israel ? ? ? ? ? ? ? ? ? 10 10Jordan ? ? ? ? ? ? ? ? ? ? 350 (7000- 10000)Kuwait ? ? ? ? ? ? ? ? ? (50) (50) (50) (50) (10) 2000Lebanon No existing or planned standards. (5,005,000)Libya ? ? ? ? ? ? ? ? ? ? (1000) 1500Morocco ? ? ? ? ? ? ? ? ? 50 50 50Oman ? ? ? ? ? ? ? ? (50) (50) (50) 500Palestinian territories No existing or planned standards. (10000)Qatar ? ? ? ? ? ? ? ? ? ? 500 500 10 500Saudia Arabia ? ? ? ? ? ? ? ? ? ? ? ? ? ? 50 5000Syria ? ? ? ? ? ? ? ? ? (50) (50) (50) (50) (50) (50) 50 7000Tunisia ? ? ? ? ? ? 10000 350 350 10 5000United Arab Emirates ? ? ? ? ? ? ? ? ? (50)/ 500 350Yemen No existing or planned standards. 10000

Sub-Saharan AfricaAngola ? ? ? ? ? ? 3000 Benin No existing standards. 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 (10000) Botswana No existing or planned standards. (500) Burkina Faso No existing standards. 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 (5000) Burundi No existing or planned standards. (5000)

Cameroon ? ? ? ? ? ? ? ? ? ? 5000 5000
292

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Report to Congress on Black Carbon

Region Country Year Current Maximum Level 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 2018 2019 2020

Sub-Saharan Africa

Cape Verde No existing or planned standards. (3000)Central African Republic No existing or planned standards. (3000- 5000)Chad No existing standards. 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 5000Côte d’Ivoire ? ? ? ? ? ? ? ? ? ? 5000 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 5000Dem. Rep. of Congo 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 (5000)Djibouti ? ? ? ? ? ? ? ? ? ? 5000 5000Equitorial Guinea No existing or planned standards. (5000- 8000)Eritrea ? ? ? ? ? ? ? ? ? ? 7000 7000Ethiopia ? ? ? ? ? ? ? ? ? ? (5000) 10000Gabon No existing or planned standards. (8000)Gambia No existing or planned standards. Ghana ? ? ? ? ? ? ? ? ? ? 5000 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 5000Guinea No existing standards. 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 (5000)Guinea Bissau No existing or planned standards. (5000)Kenya ? ? ? ? ? ? ? ? (5000) 10000Liberia No existing standards. 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 (5000)Madagascar ? ? ? ? ? ? ? ? ? ? 5000 500Malawi ? ? ? ? ? ? ? ? 500 500Mali No existing standards. 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 (10000)Mauritania No existing or planned standards. (5000)Mauritius ? 2500 2500Mozambique ? ? ? ? ? ? ? ? 500 500Namibia No existing or planned standards. (500)Niger ? ? ? ? ? ? ? ? ? ? 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 10000Nigeria ? ? ? ? ? ? ? ? ? ? (1330) 5000Republic of the Congo ? ? ? ? ? ? ? ? ? ? (1000) 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 10000Senegal ? ? ? ? ? ? ? ? ? ? 5000Sierra Leone No existing standards. 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 (5000)South Africa ? ? ? ? ? ? 500 -50 500Tanzania ? ? ? ? ? ? ? ? ? ? 5000 5000Togo ? ? ? ? ? ? ? ? ? ? 10000 3500 3500 3500 3500 3500 3500 3500 3500 3500 50 10000Uganda ? ? ? ? ? ? ? ? ? ? 5000 5000
Efforts to Limit Diesel Fuel Sulfur Levels

Report to Congress on Black Carbon 293

Region Country Year Current Maximum Level 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 2018 2019 2020

Zambia ? ? ? ? ? ? ? ? ? ? 7500 7500Zimbabwe ? ? ? ? ? ? ? ? 500

Notes: 

1. Parentheses indicates that fuel of a given sulfur level is available or sold in that country, though the national standard may differ. 2. Strikethrough numbers indicate that fuel with sulfur in excess of the local standard is commonly sold.3. Gray-shaded numbers indicate intergovernmental agreement on future standards.4. Underlined numbers refer to agreements made by national ministerial-level officials.5. Italicized numbers refer to agreements made by national officials below ministerial level.

Source Material:

1. Partnership for Clean Fuels and Vehicles, United Nations Environmental Programme (http://www.unep.org/pcfv).2. Krylov et al. (2005) Current problems. Low-sulfur diesel fuels: pluses and minuses. Chemistry and Technology of Fuels and Oils 41: 423-428.3. 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.
294

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Table A4-2. International Fuel Sulfur Limits for C3 Marine Fuel, by Target Year. (Source: U.S. EPA)

Global ECA

2004 45,000 ppm 2005 15,000 ppm

2012 35,000 ppm 2010 10,000 ppm

2020 5,000 ppm 2015 1,000 ppm
U.S. Emission Standards for Mobile 

Sources

Appendix 5

Table A5-1. On-Road Regulations for PM Control. (Source: U.S. EPA)

Regulation Sector Model Years 

Applicable

PM 

Standard Units

Tier 0 rule light duty diesel trucks 1 1981-1993 0.26 g/mile

Tier 0 rule light duty diesel trucks 2 1981-1993 0.13 g/mile

1985 Heavy Duty Diesel Rule heavy-duty highway CI engines 1988-90 0.60 g/bhp-hr

1991 Heavy Duty Diesel Rule heavy-duty highway CI engines 1991-1993 0.25 g/bhp-hr

1991 Heavy Duty Diesel Rule urban buses 1991-1993 0.10 g/bhp-hr

Tier 1 rule light duty diesel trucks 1 and 2 1994-1999 0.10 g/mile

1994 Heavy Duty Diesel Rule heavy-duty highway CI engines 1994-2006 0.10 g/bhp-hr

1994 Heavy Duty Diesel Rule urban buses 1994-1995 0.07 g/bhp-hr

1994 Heavy Duty Diesel Rule urban buses 1996-2006 0.05 g/bhp-hr

2001 Heavy Duty Diesel Rule heavy duty onroad CI engines 2007+ 0.01 g/bhp-hr

NLEV rule light duty diesel LEV cars and trucks 1999-2003 0.08 g/mile

NLEV rule light duty diesel ZLEV cars and trucks 1999-2003 0.04 g/mile

2000 Tier 2 Rule LDV LLDT bins 2-6 2004 0.01 g/mile

2000 Tier 2 Rule LDV LLDT - 50% bins 2-6 2005 0.01 g/mile

2000 Tier 2 Rule LDV LLDT - 75% bins 2-6 2006 0.01 g/mile

2000 Tier 2 Rule LDV LLDT - bins 2-6 2007 0.01 g/mile

2000 Tier 2 Rule LDV LLDT - 25% bins 7-8 2004 0.02 g/mile

2000 Tier 2 Rule LDV LLDT - 50% bins 7-8 2005 0.02 g/mile

2000 Tier 2 Rule LDV LLDT - 75% bins 7-8 2006 0.02 g/mile

2000 Tier 2 Rule LDV LLDT - bins 7-8 2007 0.02 g/mile

2000 Tier 2 Rule LDV LLDT - bin 9 2004-2006 0.06 g/mile

2000 Tier 2 Rule LDV LLDT - bin 10 2004-2006 0.08 g/mile

Report to Congress on Black Carbon 295
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Table A5-2. On-Road Gasoline-Vehicle Regulations. (Source: U.S. EPA)

Regulation Sector Model Years 

Applicable

PM 

Standard Units

2000 Tier 2 Rule Onroad Gasoline - 25% bins 2-6 2004 0.01 g/mile

2000 Tier 2 Rule Onroad Gasoline - 50% bins 2-6 2005 0.01 g/mile

2000 Tier 2 Rule Onroad Gasoline - 75% bins 2-6 2006 0.01 g/mile

2000 Tier 2 Rule Onroad Gasoline - bins 2-6 2007 0.01 g/mile

2000 Tier 2 Rule Onroad Gasoline - 25% bins 7-8 2004 0.02 g/mile

2000 Tier 2 Rule Onroad Gasoline - 50% bins 7-8 2005 0.02 g/mile

2000 Tier 2 Rule Onroad Gasoline - 75% bins 7-8 2006 0.02 g/mile

2000 Tier 2 Rule Onroad Gasoline - bins 7-8 2007 0.02 g/mile

2000 Tier 2 Rule Onroad Gasoline - bin 9 2004-2006 0.06 g/mile

2000 Tier 2 Rule Onroad Gasoline - bin 10 2004-2006 0.08 g/mile

2001 Heavy Duty Diesel Rule Heavy Duty Onroad Gasoline 50% 2008 0.01 g/bhp-hr

2001 Heavy Duty Diesel Rule Heavy Duty Onroad Gasoline 100% 2009 0.01 g/bhp-hr

Table A5-3. Nonroad Diesel Regulations. (Source: U.S. EPA)

Regulation Sector Model Years 

Applicable

PM 

Standard Units

2004 Nonroad Tier 4 Nonroad Diesel, hp<25 2008 0.3 g/bhp-hr

2004 Nonroad Tier 4 Nonroad Diesel, 25≤hp<75 2008 0.22 g/bhp-hr

2004 Nonroad Tier 4 Nonroad Diesel, 75≤hp<175 2012 0.01 g/bhp-hr

2004 Nonroad Tier 4 Nonroad Diesel, 175≤hp<750 2011 0.01 g/bhp-hr

2004 Nonroad Tier 4 Nonroad Diesel, hp750 2011 0.075 g/bhp-hr

2004 Nonroad Tier 4 Nonroad Diesel, hp750 generator sets 2015 0.02 g/bhp-hr

2004 Nonroad Tier 4 Nonroad Diesel, hp750, all other eqpt 2015 0.03 g/bhp-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, kW<8 2000 1 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, kW<8 2005 0.8 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, 8≤kW<19 2000 0.8 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, 8≤kW<19 2005 0.8 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, 19≤kW<38 1999 0.8 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, 19≤kW<38 2004 0.6 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, 37≤kW<75 2004 0.4 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, 75≤kW<130 2003 0.3 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, 130≤kW<225 1996 0.54 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, 130≤kW<225 2003 0.2 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, 225≤kW<450 1996 0.54 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, 225≤kW<450 2001 0.2 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, 450≤kW<560 1996 0.54 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, 450≤kW<560 2002 0.2 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, kW>560 2000 0.54 g/kW-hr

1998 Nonroad Diesel Engines Rule Nonroad Diesel, kW>560 2006 0.2 g/kW-hr
U.S. Emission Standards for Mobile Sources

Report to Congress on Black Carbon 297

Table A5-4. Locomotive Regulations.a (Source: U.S. EPA)

Regulation Sector Model Years 

Applicable

PM 

Standard Units

2008 LocoMarine Rule Locomotive - Line Haul Tier 0 (1973-1992) 0.22 g/bhp-hr

2008 LocoMarine Rule Locomotive - Switch Tier 0 (1973-2001) 0.26 g/bhp-hr

2008 LocoMarine Rule Locomotive - Line Haul Tier 1 (1993-2004) 0.22 g/bhp-hr

2008 LocoMarine Rule Locomotive - Switch Tier 1 (2002-2004) 0.26 g/bhp-hr

2008 LocoMarine Rule Locomotive - Line Haul Tier 2 (2005-2011) 0.10 g/bhp-hr

2008 LocoMarine Rule Locomotive - Switch Tier 2 (2005-2010) 0.13 g/bhp-hr

2008 LocoMarine Rule Locomotive - Line Haul Tier 3 (2012-2014) 0.10 g/bhp-hr

2008 LocoMarine Rule Locomotive - Switch Tier 3 (2011-2014) 0.10 g/bhp-hr

2008 LocoMarine Rule Locomotive - Line Haul Tier 4 (2015+) 0.03 g/bhp-hr

2008 LocoMarine Rule Locomotive - Switch Tier 4 (2015+) 0.03 g/bhp-hr

1997 Locomotive Stds Locomotive - Line Haul Tier 0 (1973-2001) 0.60 g/bhp-hr

1997 Locomotive Stds Locomotive - Switch Tier 0 (1973-2001) 0.72 g/bhp-hr

1997 Locomotive Stds Locomotive - Line Haul Tier 1 (2002-2004) 0.45 g/bhp-hr

1997 Locomotive Stds Locomotive - Switch Tier 1 (2002-2004) 0.54 g/bhp-hr

1997 Locomotive Stds Locomotive - Line Haul Tier 2 (2005+) 0.20 g/bhp-hr

1997 Locomotive Stds Locomotive - Switch Tier 2 (2005+) 0.24 g/bhp-hr

a 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.
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Table A5-5. Commercial Marine Regulations. (Source: U.S. EPA)

Regulation Sector Model Years 

Applicable

PM 

standard Units

2008 LocoMarine Rule - Tier 3 C1 Commercial Stnd Power Density, <19kW Max power, 

<0.9L/cylinder 2009+ 0.30 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Commercial Stnd Power Density, 19<75kW Max 

power, <0.9L/cylinder 2009+ 0.22 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Commercial Stnd Power Density, 75<3700kW Max 

power, <0.9L/cylinder 2012 0.10 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Commercial Stnd Power Density, 75<3700kW Max 

power, 0.9<1.2L/cylinder 2013 0.09 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Commercial Stnd Power Density, 75<3700kW Max 

power, 1.2<2.5L/cylinder 2014 0.08 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Commercial Stnd Power Density, 75<3700kW Max 

power, 2.5<3.5L/cylinder 2013 0.08 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Commercial Stnd Power Density, 75<3700kW Max 

power, 3.5<7.0L/cylinder 2012 0.08 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Rec and Com High Power Density, <19kW Max power, 

<0.9L/cylinder 2009+ 0.30 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Rec and Com High Power Density, 19<75kW Max 

power, <0.9L/cylinder 2009+ 0.22 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Rec and Com High Power Density, 75<3700kW Max 

power, <0.9L/cylinder 2012 0.11 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Rec and Com High Power Density, 75<3700kW Max 

power, 0.9<1.2L/cylinder 2013 0.10 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Rec and Com High Power Density, 75<3700kW Max 

power, 1.2<2.5L/cylinder 2014 0.09 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Rec and Com High Power Density, 75<3700kW Max 

power, 2.5<3.5L/cylinder 2013 0.09 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C1 Rec and Com High Power Density, 75<3700kW Max 

power, 3.5<7.0L/cylinder 2012 0.08 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C2,<3700kW,7-<15L/cylinder 2013 0.10 g/bhp-hr

2008 LocoMarine Rule - Tier 3 C2,<3700kW,15-<30L/cylinder 2014 0.20 g/bhp-hr

2008 LocoMarine Rule - Tier 4 C1&C2, >3700 kW 2014 0.09 g/bhp-hr

2008 LocoMarine Rule - Tier 4 C1&C2, >3700 kW 2016 0.04 g/bhp-hr

2008 LocoMarine Rule - Tier 4 C1&C2, 600 to <3700 2014 0.03 g/bhp-hr

1999 C1&C2 Marine Engine Rule C1, power ≤37 kW disp. <0.9 2005+ 0.40 g/kW-hr

1999 C1&C2 Marine Engine Rule C1, 0.9≤disp.< 1.2 2004+ 0.30 g/kW-hr

1999 C1&C2 Marine Engine Rule C1, 1.2≤disp.<2.5 2004+ 0.20 g/kW-hr

1999 C1&C2 Marine Engine Rule C1, 2.5≤disp.<5.0 2007+ 0.20 g/kW-hr

1999 C1&C2 Marine Engine Rule C2, 5.0≤disp.<15.0 2007+ 0.27 g/kW-hr

1999 C1&C2 Marine Engine Rule C2, 15.0≤disp.<30.0 2007+ 0.50 g/kW-hr
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 heavyduty 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 

Appendix 6

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 lightduty 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 technologybased 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
300

Appendix 6

Report to Congress on Black Carbon

Figure A6-1. Heavy-Duty Highway Diesel PM Emissions Standards in the Americas and the Caribbean 

(Logarithmic Scale). (Source: U.S. EPA)

Figure A6-2. Heavy-Duty Highway Diesel PM Emissions Standards in Asia and Australia (Logarithmic 

Scale). (Source: U.S. EPA)
International Emission Standards for Heavy-Duty Vehicles

Report to Congress on Black Carbon 301

Figure A6-3. Heavy-Duty Highway Diesel PM Emissions Standards in Europe (Logarithmic Scale). 

(Source: U.S. EPA)

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.

y It is important to develop standardized definitions 

of BC, BrC, non-light-absorbing OC and other 

light-absorbing aerosols (2.3). 

y It is also important to establish corresponding 

reference materials for needed particle properties 

(2.3, 5.2). 

y Interpretation and harmonization of existing 

ambient and emissions measurements are 

Appendix 7

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. 

y 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 nonlight-absorbing (scattering) portions of OC are 

especially important to improve inventories, to 

improve model estimates of SOA, and to evaluate 

models.

y 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. 

y 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 

Report to Congress on Black Carbon 303
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.

y 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.

y 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.

y 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 cobenefits 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 

304

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Report to Congress on Black Carbon
emissions side and on the “usage” (activity level) side 

as well on determining how best to allocate these 

emissions both spatially and temporally:

y 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. 

y Emissions estimates for nonroad applications 

need to be developed and collated separately 

from on-road operations (4.4.1). 

y In the case of biomass burning, all fires need to be 

better captured for their emissions characteristics 

and activity level. 

y 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). 

y 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 fieldbased 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.

y Steps need to be taken to better engage 

international scientists, governments, and 

regulatory agencies in collaborations of 

technology assessments to help improve baseyear 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.

y 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. 

y 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 newertechnology 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 

Research Needs

Report to Congress on Black Carbon 305
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.

y Quantitative measures for describing uncertainties 

in emissions should be advanced, as was 

suggested in Section 4.4.4. 

y As a first step, a qualitative description of 

uncertainty in emissions estimates, sectorby-sector, for domestic emissions inventories 

including models should be performed. 

y The uncertainty that stems from combining BC 

data sets collected by several measurement 

techniques (for regional and global inventories) 

should be estimated. 

y 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. 

y 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).

y 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. 

y 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. 

y 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). 

y 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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Report to Congress on Black Carbon
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. 

y 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.

y 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, Andes, 

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-relevant 

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:

y 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. 

y 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. longlived 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 

Research Needs

Report to Congress on Black Carbon 307
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.

y 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. 

y 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 costeffective 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 

308

Appendix 7

Report to Congress on Black Carbon
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, 

Research Needs

Report to Congress on Black Carbon 309
beginning with a synthesis of available research 

results (11.5.1, 11.6) would be especially useful in 

considering mitigation options. This type of analysis 

could conceivably include (1) total life-cycle studies 

of mitigation methods, both individually and in 

combination with others, for ecosystems impacts, 

(2) total life-cycle studies of the net climate forcing 

impacts arising from the use of these methods, or 

(3) total life-cycle studies of the economic impacts 

of the use of these methods.

310

Appendix 7

Report to Congress on Black Carbon
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Research Triangle Park, NC

Office of Air Quality Planning and Standards Office of Air and Radiation

Protection AgencyEnvironmental United States Publication No. EPA-450/R-12-001

March 2012