United States

Environmental Protection

Agency

Air

EPA-452/R-97-007

December 1997

Mercury Study

Report to Congress

c7o032-1-1

Office of Air Quality Planning & Standards

and

Office of Research and Development

Volume V:

Health Effects of Mercury

and Mercury Compounds
MERCURY STUDY REPORT TO CONGRESS

VOLUME V:

HEALTH EFFECTS OF MERCURY AND MERCURY COMPOUNDS

December 1997

Office of Air Quality Planning and Standards

and

Office of Research and Development

U.S. Environmental Protection Agency
i

TABLE OF CONTENTS

Page

U.S. EPA AUTHORS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . iv

SCIENTIFIC PEER REVIEWERS ...................................................... v

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

LIST OF TABLES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ix

LIST OF FIGURES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xii

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

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

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

2. TOXICOKINETICS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-1

2.1 Absorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-1

2.1.1 Elemental Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-1

2.1.2 Inorganic Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-2

2.1.3 Methylmercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-3

2.2 Distribution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-4

2.2.1 Elemental Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-4

2.2.2 Inorganic Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-4

2.2.3 Methylmercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-5

2.3 Metabolism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-6

2.3.1 Elemental Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-6

2.3.2 Inorganic Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-6

2.3.3 Methylmercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-6

2.4 Excretion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-7

2.4.1 Elemental Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-7

2.4.2 Inorganic Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-8

2.4.3 Methylmercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-8

2.5 Biological Monitoring . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-9

2.5.1 Elemental Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-9

2.5.2 Inorganic Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-11

2.5.3 Methylmercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-11

2.5.4 Methods of Analysis for Measuring Mercury in Biological Samples . . . . . . 2-12

2.6 Studies on Pharmacokinetic Models . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-12

2.6.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-12

2.6.2 Inorganic mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-13

2.6.3 Methylmercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-13

2.6.4 Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-15

3. BIOLOGICAL EFFECTS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-1

3.1 Elemental Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-1

3.1.1 Critical Noncancer Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-1

3.1.2 Cancer Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-5

3.1.3 Other Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-9
TABLE OF CONTENTS (continued)

Page

ii

3.2 Inorganic Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-35

3.2.1 Critical Noncancer Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-35

3.2.2 Cancer Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-39

3.2.3 Other Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-43

3.3 Methylmercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-59

3.3.1 Critical Noncancer Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-59

3.3.2 Cancer Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-66

3.3.3 Other Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-73

4. SUSCEPTIBLE POPULATIONS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-1

5. INTERACTIONS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-1

6. HAZARD IDENTIFICATION AND DOSE-RESPONSE ASSESSMENT . . . . . . . . . . . . . . . 6-1

6.1 Background . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-1

6.1.1 Hazard Identification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-1

6.1.2 Dose-response Assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-4

6.2 Hazard Identification for Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-7

6.2.1 Developmental Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-8

6.2.2 Germ Cell Mutagenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-11

6.2.3 Carcinogenic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-13

6.3 Dose-Response Assessment For Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-16

6.3.1 Systemic Noncancer Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-16

6.3.2 Developmental Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-54

6.3.3 Germ Cell Mutagenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-55

6.3.4 Carcinogenic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-55

6.4 Risk Assessments Done By Other Groups . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-56

6.4.1 Food and Drug Administration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-56

6.4.2 ATSDR . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-57

6.4.3 Department of Energy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-57

6.4.4 National Institute of Environmental Health Sciences (NIEHS) . . . . . . . . . . . 6-58

6.4.5 Department of Labor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-58

6.4.6 Various States . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-58

6.4.7 World Health Organization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-58

6.4.8 ACGIH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-59

7. ONGOING RESEARCH AND RESEARCH NEEDS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-1

7.1 Ongoing Research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-1

7.2 Research Needs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-3

8. REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8-1
TABLE OF CONTENTS (continued)

Page

iii

APPENDIX A DOSE CONVERSIONS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A-1

APPENDIX B SUMMARIES FOR THE INTEGRATED RISK 

INFORMATION SYSTEM (IRIS) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B-1

APPENDIX C ATTENDEES OF U.S. EPA PEER REVIEW WORKSHOP

ON MERCURY ISSUES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . C-1

APPENDIX D UNCERTAINTY ANALYSIS OF THE METHYLMERCURY RfD . . . . . . . . . . . . D-1
iv

U.S. EPA AUTHORS

Principal Authors:

Beth Hassett-Sipple Rita Schoeny, Ph.D.

Office of Air Quality Office of Water

 Planning and Standards Washington, DC

Research Triangle Park, NC

Jeff Swartout

National Center for Environmental

 Assessment-Cincinnati

Office of Research and Development

Cincinnati, OH

Contributing Authors:

Kathryn R. Mahaffey, Ph.D.

National Center for Environmental

 Assessment-Washington

Office of Research and Development

Washington, DC

Glenn E. Rice

National Center for Environmental

 Assessment-Cincinnati

Office of Research and Development

Cincinnati, OH
v

SCIENTIFIC PEER REVIEWERS

Dr. William J. Adams* U.S. Department of Energy

Kennecott Utah Corporation Policy Office, Washington D.C.

Dr. Brian J. Allee Dr. Rick Canady

Harza Northwest, Incorporated Agency for Toxic Substances and Disease

Dr. Thomas D. Atkeson

Florida Department of Environmental Dr. Rufus Chaney

Protection U.S. Department of Agriculture

Dr. Donald G. Barnes* Dr. Joan Daisey*

U.S. EPA Science Advisory Board Lawrence Berkeley National Laboratory

Dr. Steven M. Bartell Dr. John A. Dellinger*

SENES Oak Ridge, Inc. Medical College of Wisconsin

Dr. David Bellinger* Dr. Kim N. Dietrich*

Children’s Hospital, Boston University of Cincinnati

Dr. Nicolas Bloom* Dr. Tim Eder

Frontier Geosciences, Inc. Great Lakes Natural Resource Center

Dr. Mike Bolger States of Michigan and Ohio

U.S. Food and Drug Administration

Dr. Peter Botros Michigan State University

U.S. Department of Energy

Federal Energy Technology Center Dr. William F. Fitzgerald

Thomas D. Brown Avery Point

U.S. Department of Energy

Federal Energy Technology Center A. Robert Flaak*

Dr. Dallas Burtraw*

Resources for the Future Dr. Bruce A. Fowler*

Dr. Thomas Burbacher*

University of Washington Dr. Katherine Flegal

Seattle National Center for Health Statistics

Dr. James P. Butler Dr. Steven G. Gilbert*

University of Chicago Biosupport, Inc.

Argonne National Laboratory

Elizabeth Campbell The Academy of Natural Sciences

Registry

National Wildlife Federation for the

Dr. Lawrence J. Fischer*

University of Connecticut

U.S. EPA Science Advisory Board

University of Maryland at Baltimore

Dr. Cynthia C. Gilmour*
SCIENTIFIC PEER REVIEWERS (continued)

vi

Dr. Robert Goyer Dr. Genevieve M. Matanoski*

National Institute of Environmental Health The Johns Hopkins University

Sciences

Dr. George Gray University of California

Harvard School of Public Health Berkeley

Dr. Terry Haines Dr. Malcolm Meaburn

National Biological Service National Oceanic and Atmospheric

Dr. Gary Heinz* U.S. Department of Commerce

Patuxent Wildlife Research Center

Joann L. Held Wisconsin Department of Natural Resources

New Jersey Department of Environmental

Protection & Energy Dr. Maria Morandi*

Dr. Robert E. Hueter*

Mote Marine Laboratory Dr. Paul Mushak

Dr. Harold E. B. Humphrey*

Michigan Department of Community Health Harvey Ness

Dr. James P. Hurley* Federal Energy Technology Center

University of Wisconsin

Madison Dr. Christopher Newland*

Dr. Joseph L. Jacobson*

Wayne State University Dr. Jerome O. Nriagu*

Dr. Gerald J. Keeler Ann Arbor

University of Michigan

Ann Arbor William O’Dowd

Dr. Ronald J. Kendall* Federal Energy Technology Center

Clemson University

Dr. Lynda P. Knobeloch* University of Florida

Wisconsin Division of Health Gainesville

Dr. Leonard Levin Dr. Jozef M. Pacyna

Electric Power Research Institute Norwegian Institute for Air Research

Dr. Steven E. Lindberg* Dr. Ruth Patterson

Oak Ridge National Laboratory Cancer Prevention Research Program

Dr. Thomas McKone*

Administration

Dr. Michael W. Meyer*

University of Texas Science Center at Houston

PB Associates

U.S. Department of Energy

Auburn University

The University of Michigan

U.S. Department of Energy

Dr. W. Steven Otwell*

Fred Gutchinson Cancer Research Center
SCIENTIFIC PEER REVIEWERS (continued)

vii

Dr. Donald Porcella Dr. Alan H. Stern

Electric Power Research Institute New Jersey Department of Environmental

Dr. Deborah C. Rice*

Toxicology Research Center Dr. David G. Strimaitis*

Samuel R. Rondberg*

U.S. EPA Science Advisory Board Dr. Edward B. Swain

Charles Schmidt

U.S. Department of Energy Dr. Valerie Thomas*

Dr. Pamela Shubat

Minnesota Department of Health Dr. M. Anthony Verity

Dr. Ellen K. Silbergeld* Los Angeles

University of Maryland

Baltimore

Dr. Howard A. Simonin*

NYSDEC Aquatic Toxicant Research Unit

Dennis Smith

U.S. Department of Energy

Federal Energy Technology Center

Dr. Ann Spacie*

Purdue University

Protection & Energy

Earth Tech

Minnesota Pollution Control Agency

Princeton University

University of California

*With EPA’s Science Advisory Board, Mercury Review Subcommitte
viii

WORK GROUP AND U.S. EPA/ORD REVIEWERS

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

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

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

Angela Bandemehr, U.S. EPA

Region 5 Eletha Brady-Roberts

Jim Darr, U.S. EPA Cincinnati, OH

Office of Pollution Prevention and Toxic

Substances Annie M. Jarabek

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

Department of Environmental Conservation

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

Office of Water Washington, DC

David Layland, U.S. EPA Susan Braen Norton

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

Karen Levy, U.S. EPA

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

Steve Levy, U.S. EPA Cincinnati, OH

Office of Solid Waste and Emergency Response

Lorraine Randecker, U.S. EPA

Office of Pollution Prevention and Toxic

Substances

Joy Taylor, State of Michigan

Department of Natural Resources

Washington, DC

National Center for Environmental Assessment

National Center for Environmental Assessment

Matthew Lorber

Washington, DC

National Center for Environmental Assessment
ix

LIST OF TABLES

Page

ES-1 Summary of U.S. EPA Hazard Identification/Dose-response

Assessment for Mercury and Mercury Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . ES-6

2-1 Reference Values for Total Mercury Concentrations in Biological Media for the General

Population . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-10

2-2 Analytical Methods for the Detection of Mercury in Biological Samples . . . . . . . . . . . 2-12

3-1 Carcinogenic Effects of Elemental Mercury in Humans: Epidemiological Studies . . . . 3-7

3-2 Lethality of Elemental Mercury in Humans: Case Studies . . . . . . . . . . . . . . . . . . . . . . . 3-9

3-3 Lethality of Elemental Mercury in Animals: Inhalation Exposure . . . . . . . . . . . . . . . . 3-10

3-4 Neurotoxicity of Elemental Mercury in Humans: Case Studies . . . . . . . . . . . . . . . . . . 3-11

3-5 Neurotoxicity of Elemental Mercury in Humans: Epidemiological Studies . . . . . . . . . 3-13

3-6 Neurotoxicity of Elemental Mercury in Animals: Inhalation Exposure . . . . . . . . . . . . . 3-15

3-7 Renal Toxicity of Elemental Mercury in Humans: Case Studies . . . . . . . . . . . . . . . . . 3-16

3-8 Renal Toxicity of Elemental Mercury in Humans: Epidemiological Studies . . . . . . . . 3-17

3-9 Renal Toxicity of Elemental Mercury in Animals: Inhalation Exposure . . . . . . . . . . . 3-18

3-10 Respiratory Toxicity of Elemental Mercury in Humans: Case Studies . . . . . . . . . . . . . 3-18

3-11 Respiratory Toxicity of Elemental Mercury in Animals: Inhalation Exposure . . . . . . . 3-20

3-12 Cardiovascular Toxicity of Elemental Mercury in Humans: Case Studies . . . . . . . . . . 3-21

3-13 Cardiovascular Toxicity of Elemental Mercury in Humans: Epidemiological

Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-21

3-14 Cardiovascular Toxicity of Elemental Mercury in Animals: Inhalation

Exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-22

3-15 Gastrointestinal Toxicity of Elemental Mercury in Humans: Case Studies . . . . . . . . . 3-23

3-16 Gastrointestinal Toxicity of Elemental Mercury in Animals: Inhalation

Exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-24

3-17 Hepatic Toxicity of Elemental Mercury in Humans: Case Study . . . . . . . . . . . . . . . . . 3-24

3-18 Hepatic Toxicity of Elemental Mercury in Animals: Inhalation Exposure . . . . . . . . . . 3-25

3-19 Hematological Toxicity of Elemental Mercury in Humans: Case Studies . . . . . . . . . . 3-25

3-20 Hematological Toxicity of Elemental Mercury in Humans: Epidemiological

Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-26

3-21 Immunotoxicity of Elemental Mercury in Humans: Case Study . . . . . . . . . . . . . . . . . . 3-27

3-22 Immunotoxicity of Elemental Mercury in Humans: Epidemiological Studies . . . . . . . 3-27

3-23 Immunotoxicity of Elemental Mercury in Animals: Inhalation Exposure . . . . . . . . . . 3-27

3-24 Dermal Toxicity of Elemental Mercury in Humans: Case Studies . . . . . . . . . . . . . . . . 3-28

3-25 Developmental Toxicity of Elemental Mercury in Humans: Case

Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-29

3-26 Developmental Toxicity of Elemental Mercury in Humans: Epidemiological

Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-30

3-27 Developmental Toxicity of Elemental Mercury in Animals . . . . . . . . . . . . . . . . . . . . . . 3-31

3-28 Reproductive Toxicity of Elemental Mercury in Humans: Epidemiological

Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-32

3-29 Reproductive Toxicity of Elemental Mercury in Animals . . . . . . . . . . . . . . . . . . . . . . . 3-33

3-30 Genotoxicity of Elemental Mercury in Humans . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-34

3-31 Incidence of Selected Lesions in Rats in the NTP (1993) 2-Year Gavage Study . . . . . . 3-41

3-32 Incidence of Renal Tubule Tumors in Male Mice in the NTP (1993)
LIST OF TABLES (continued)

x

2-Year Gavage Study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-42

3-33 Carcinogenic Effects of Inorganic Mercury in Animals: Oral Exposure . . . . . . . . . . . . 3-43

3-34 Lethality of Inorganic Mercury in Humans: Case Study . . . . . . . . . . . . . . . . . . . . . . . . 3-43

3-35 Lethality of Inorganic Mercury in Animals: Oral Exposure . . . . . . . . . . . . . . . . . . . . . 3-44

3-36 Neurotoxicity of Inorganic Mercury in Humans: Case Studies . . . . . . . . . . . . . . . . . . . 3-45

3-37 Neurotoxicity of Inorganic Mercury in Animals: Oral Exposure . . . . . . . . . . . . . . . . . 3-45

3-38 Renal Toxicity of Inorganic Mercury in Humans: Case Studies . . . . . . . . . . . . . . . . . . 3-46

3-39 Renal Toxicity of Inorganic Mercury in Animals: Oral Exposure . . . . . . . . . . . . . . . . 3-46

3-40 Renal Toxicity of Inorganic Mercury in Animals: Inhalation Exposure . . . . . . . . . . . . 3-48

3-41 Cardiovascular Toxicity of Inorganic Mercury in Animals . . . . . . . . . . . . . . . . . . . . . . 3-48

3-42 Gastrointestinal Toxicity of Inorganic Mercury in Humans: Case Studies . . . . . . . . . . 3-49

3-43 Gastrointestinal Toxicity of Inorganic Mercury in Animals . . . . . . . . . . . . . . . . . . . . . . 3-50

3-44 Hepatic Toxicity of Inorganic Mercury in Animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-51

3-45 Immunotoxicity of Inorganic Mercury in Animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-52

3-46 Developmental Toxicity of Inorganic Mercury in Animals: Inhalation

Exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-52

3-47 Developmental Toxicity of Inorganic Mercury in Animals: Oral Exposure . . . . . . . . . 3-54

3-48 Reproductive Toxicity of Inorganic Mercury in Humans: Case Study . . . . . . . . . . . . . 3-56

3-49 Reproductive Toxicity of Inorganic Mercury in Animals . . . . . . . . . . . . . . . . . . . . . . . 3-57

3-50 Genotoxicity of Inorganic Mercury in Humans . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-57

3-51 Genotoxicity of Inorganic Mercury in Animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-59

3-52 Carcinogenic Effects of Methylmercury in Humans: Epidemiological Studies . . . . . . 3-68

3-53 Carcinogenic Effects of Methylmercury in Animals: Oral Exposure . . . . . . . . . . . . . . 3-70

3-54 Lethality of Methylmercury in Humans: Case Study of Oral Exposure . . . . . . . . . . . . 3-73

3-55 Lethality of Methylmercury in Humans: Case Studies of Inhalation Exposure . . . . . . 3-74

3-56 Lethality of Methylmercury in Animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-74

3-57 Neurotoxicity of Methylmercury in Humans: Case Studies of Oral Exposure . . . . . . . 3-77

3-58 Neurotoxicity of Methylmercury in Humans: Case Studies of Inhalation

Exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-78

3-59 Neurotoxicity of Methylmercury in Animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-80

3-60 Renal Toxicity of Methylmercury in Animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-82

3-61 Cardiovascular Toxicity of Methylmercury in Humans: Case Study . . . . . . . . . . . . . . 3-82

3-62 Cardiovascular Toxicity of Methylmercury in Animals . . . . . . . . . . . . . . . . . . . . . . . . . 3-82

3-63 Gastrointestinal Toxicity of Methylmercury in Animals . . . . . . . . . . . . . . . . . . . . . . . . 3-84

3-64 Immunotoxicity of Methylmercury in Animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-84

3-65 Dermal Toxicity of Methylmercury in Humans: Epidemiological Study . . . . . . . . . . . 3-85

3-66 Developmental Toxicity of Methylmercury in Humans: Case Studies . . . . . . . . . . . . . 3-86

3-67 Developmental Toxicity of Methylmercury in Humans: Epidemiologic Studies . . . . . 3-87

3-68 Developmental Toxicity of Methylmercury in Animals . . . . . . . . . . . . . . . . . . . . . . . . . . 3-88

3-69 Reproductive Toxicity of Methylmercury in Animals . . . . . . . . . . . . . . . . . . . . . . . . . . 3-94

3-70 Genotoxicity of Methylmercury in Humans: Case Study . . . . . . . . . . . . . . . . . . . . . . . 3-95

3-71 Genotoxicity of Methylmercury in Humans: Epidemiology Study . . . . . . . . . . . . . . . . . 3-95

3-72 Genotoxicity of Methylmercury in Cats . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-96

5-1 Interactions of Mercury with Other Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-3

6-1 Consensus Decisions of Peer Review Panel . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-18

6-2 Available Data on Hair:Blood Ration (Total Hg) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-24

6-3 Incidence of Effects in Iraqi Children by Exposure Group . . . . . . . . . . . . . . . . . . . . . . 6-26
LIST OF TABLES (continued)

xi

6-4 Incidence of Effects in Iraqi Adults by Exposure Group . . . . . . . . . . . . . . . . . . . . . . . . . 6-26

6-5 Methylmercury Benchmark Dose Estimates (ppm hair) . . . . . . . . . . . . . . . . . . . . . . . . . 6-28

6-6 Methylmercury Benchmark Dose Estimates (ppb blood) . . . . . . . . . . . . . . . . . . . . . . . . . 6-28

6-7 Density-Based Dose Groupings . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-29

6-8 Uniform Dose Groupings . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-30

6-9 Benchmark Dosed Calculated on Data from Marsh et al. (1987) . . . . . . . . . . . . . . . . . . . 6-30

6-10 Summary of Benchmark Doses (BMD) Estimated for Methylmercury . . . . . . . . . . . . . . 6-32

6-11 Results of Revised Denver Developmental Screening Test (DDST) Administered

to Seychellois Children in Cross-sectional Study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-33

6-12 Estimates of No Observed Adverse Effect Levels (NOAELs) and Lowest Observed

Adverse Effect Levels (LOAELs) from Human Studies . . . . . . . . . . . . . . . . . . . . . . . . . 6-34

6-13 Estimates of NOAELs and LOAELs from Animal Studies . . . . . . . . . . . . . . . . . . . . . . . 6-39

6-14 NOAELs and LOAELs for Developmental Toxicity of Methylmercury in Animals . . . . 6-42

6-15 Studies in Non-Human Primates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-46

7-1 Ongoing Research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-1
xii

LIST OF FIGURES

Page

6-1 Density of Data Points Relative to Hg Concentration in Hair

for Iraqi Cohort Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-29
xiii

LIST OF SYMBOLS, UNITS AND ACRONYMS

ATSDR Agency for Toxic Substances and Disease Registry

BML Biological monitoring level

bw Body weight

CAA Clean Air Act as amended in 1990

CHO Chinese hamster ovary

C.I. Confidence interval

CNS Central nervous system

CRAVE Carcinogen Risk Assessment Verification Endeavor

DDST Denver Developmental Screen Test

DHHS Department of Health and Human Services

DNA Deoxyribonucleic acid

DWEL Drinking water equivalent level

ECG Electrocardiogram

EEG Electroencephalogram

EPA Environmental Protection Agency

FDA Food and Drug Administration

GABA Gamma aminobutyric acid

Gd Gestation day

HEC Human equivalent concentration

Hg Mercury

HgU Urinary mercury

IgG Immunoglobulin G

IRIS Integrated Risk Information System

LC Lethal concentration killing 50 percent of the animals tested (inhalation) 50

LD Lethal dose killin 50 g 50 percent of the animals tested

LOAEL Lowest-observed-adverse-effect level

MF Modifying factor

MMAD Mass median aerodynamic diameter

MMC Methylmercuric chloride

MMH Methylmercuric hydroxide

MRL Minimal risk level

MTD Maximum tolerated dose

NAG N-acetyl-b-glycosaminidase

NADH Reduced nicotinamide adenine dinucleotide

NADPH Reduced nicotinamide adenine dinucleotide phosphate

NOAEL No-observed-adverse-effect level

NS Not specified

NTP National Toxicology Program

PMA Phenyl mercuric acetate

ppd Postpartum day

RfD Reference dose (oral)

RfD Reference dose for develo DT pmental toxicity

RfC Reference concentration (inhalation)

SCE Sister chromatid exchange

SGPT Serum glutamic-pyruvic transaminase

SH Sulfhydryl groups
LIST OF SYMBOLS, UNITS AND ACRONYMS (continued)

xiv

SMR Standard mortality ratio

TOLD Test of Language Development

TWA Time-weighted average

UF Uncertainty factor

UF Uncertaint A y factor for interspecies extrapolation

UF Uncertaint H y factor for intraspecies extrapolation (animal to human)

UF Uncertaint L y factor for use of a LOAEL

UF Uncertaint S y factor for use of a subchronic-duration study

WHO World Health Organization
ES-1

EXECUTIVE SUMMARY

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

Environmental Protection Agency (U.S. EPA) to submit to Congress a comprehensive study on

atmospheric emissions of mercury. This document, which covers the human health effects of mercury

and mercury compounds, is one volume of U.S. EPA's eight-volume Report in response to this directive.

Mercury is a naturally occurring element that is found in air, water and soil. It exists in any of

three valence states: Hg (elemental mercury), Hg (mercurous mercury), or Hg (mercuric mercury). 0 2+ 2+

2

Most of the population of the earth have some exposure to mercury as a result of normal daily activities. 

The general population may be exposed to mercury through inhalation of ambient air; consumption of

contaminated food, water, or soil; and/or dermal exposure to substances containing mercury. In addition,

some quantity of mercury is released from dental amalgam.

The health effects literature contains many investigations of populations with potentially high

exposure to mercury, including industrial workers, people living near point sources of mercury

emissions, people who consume large amounts of fish, and dental professionals. There also are

numerous studies of populations unintentionally exposed to high levels of mercury, such as the Minamata

poisoning episode in Japan. Volume IV (An Assessment Exposure to Mercury in the United States)

presents measured and predicted mercury exposure for various U.S. populations.

The purpose of this volume, Volume V, is to summarize the available health effects information

for mercury and mercury compounds and to present U.S. EPA's analysis for two critical pieces of the risk

assessment paradigm described by the National Academy of Sciences in 1983. Specifically, this volume

contains the hazard identification and dose-response assessments for three forms of mercury: elemental

mercury, mercuric chloride (inorganic mercury),and methylmercury (organic mercury). In order to

characterize risk for any populations, the evaluations presented in this volume must be combined with the

assessment of exposure presented in Volume IV.

Volume V is not intended to be an exhaustive survey of the voluminous health effects literature

available for mercury. Rather, the purpose is to present a brief survey of the studies relevant for

assessing potential human health effects and to present more detailed information on those studies which

form the basis for U.S. EPA's hazard identification and dose-response assessments. The three forms of

mercury which are emphasized in this volume were selected based on data indicating that these are the

predominant forms of mercury to which humans are exposed. In addition, examination of the published

literature indicates that most health data are on these forms. It is acknowledged that certain populations

can be exposed to many types of organic mercurials, such as antiseptics and pesticides. Volume V,

however, deals with methylmercury except in cases where information on another organic is presented

for illustrative purposes.

Toxicokinetics

The toxicokinetics (i.e., absorption, distribution, metabolism, and excretion) of mercury is highly

dependent on the form of mercury to which a receptor has been exposed. Below is a brief summary of

the toxicokinetics information for elemental mercury, mercuric chloride, and methylmercury. Chapter 2

contains a more complete summary of the toxicokinetics information available for mercury.
ES-2

Elemental Mercury

The absorption of elemental mercury vapor occurs rapidly through the lungs, but it is poorly

absorbed from the gastrointestinal tract. Once absorbed, elemental mercury is readily distributed

throughout the body; it crosses both placental and blood-brain barriers. Elemental mercury is oxidized to

inorganic divalent mercury by the hydrogen peroxidase-catalase pathway, which is present in most

tissues. The distribution of absorbed elemental mercury is limited primarily by the oxidation of

elemental mercury to the mercuric ion as the mercuric ion has a limited ability to cross the placental and

blood-brain barriers. Once elemental mercury crosses these barriers and is oxidized to the mercuric ion,

return to the general circulation is impeded, and mercury can be retained in brain tissue. The elimination

of elemental mercury occurs via urine, feces, exhaled air, sweat, and saliva. The pattern of excretion is

dependent on the extent to which elemental mercury has been oxidized to mercuric mercury.

Inorganic Mercury

Absorption of inorganic mercury through the gastrointestinal tract varies with the particular

mercuric salt involved. Absorption decreases with decreasing solubility. Estimates of the percentage of

inorganic mercury that is absorbed vary; as much as 20% may be absorbed. Available data indicate that

absorption of mercuric chloride from the gastrointestinal tract results from an electrostatic interaction

with the brush border membrane and limited passive diffusion. Increases in intestinal pH, high doses of

mercuric chloride causing a corrosive action, a milk diet (e.g., neonates) and increases in pinocytotic

activity in the gastrointestinal tract (e.g., neonates) have all been associated with increased absorption of

inorganic mercury. Inorganic mercury has a limited capacity for penetrating the blood-brain or placental

barriers. There is some evidence indicating that mercuric mercury in the body following oral exposures

can be reduced to elemental mercury and excreted via exhaled air. Because of the relatively poor

absorption of orally administered inorganic mercury, the majority of the ingested dose in humans is

excreted through the feces.

Methylmercury

Methylmercury is rapidly and extensively absorbed through the gastrointestinal tract. Absorption

information following inhalation exposures is limited. This form of mercury is distributed throughout the

body and easily penetrates the blood-brain and placental barriers in humans and animals. Methylmercury

transport into tissues appears to be mediated by the formation of a methylmercury-cysteine complex. 

This complex is structurally similar to methionine and is transported into cells via a widely distributed

neutral amino acid carrier protein. Methylmercury in the body is considered to be relatively stable and is

only slowly demethylated to form mercuric mercury in rats. It is hypothesized that methylmercury

metabolism may be related to a latent or silent period observed in epidemiological studies observed as a

delay in the onset of specific adverse effects. Methylmercury has a relatively long biological half-life in

humans; estimates range from 44 to 80 days. Excretion occurs via the feces, breast milk, and urine.

Biological Monitoring/Pharmacokinetic Models

Chapter 2 provides information on biological monitoring of mercury as well as a summary of the

development of pharmacokinetic models for mercury. The most common biological samples analyzed

for mercury are blood, urine, and scalp hair. The methods most frequently used to determine the mercury

levels in these sample types include atomic absorption spectrometry, neutron activation analysis, X-ray

fluorescence, and gas chromatography.
ES-3

Both simple and complex multi-compartmental models have been described in the literature. A

recent report (Gearhart et al. 1995) presents an approach based upon data from human, rat, and monkey

data that could be used for characterizing dose-response data both adults and neonates.

Biological Effects

Chapter 3 presents summary information on the toxicity of elemental mercury, mercuric mercury

and methylmercury to various organ systems. The primary targets for toxicity of mercury and mercury

compounds are the nervous system, the kidney, and the developing fetus. Other systems that may be

affected include the respiratory, cardiovascular, gastrointestinal, hematologic, immune, and reproductive

systems. For each form of mercury and each of the endpoints addressed, information from

epidemiological studies, human case studies, and animal toxicity studies is summarized in tabular form. 

Critical studies are discussed in the accompanying text.

Elemental Mercury

A number of epidemiological studies have been conducted that examined cancer mortality and/or

morbidity among workers occupationally exposed to elemental mercury. All of these studies, however,

have limitations which compromise the interpretation of their results; these limitations include small

sample sizes, probable exposure to other known lung carcinogens, failure to consider confounding factors

such as smoking, and/or failure to observe correlations between estimated exposure and the cancer

incidence. Only one animal study was identified that examined cancer incidence in animals exposed (by

injection) to elemental mercury. While tumors were found at contact sites, the study was incompletely

reported as to controls and statistics and, thus, considered inadequate for the purpose of risk assessment. 

Findings from genotoxicity assays are limited and do not provide supporting evidence for a carcinogenic

effect of elemental mercury.

Effects on the nervous system appear to be the most sensitive toxicological endpoint observed

following exposure to elemental mercury. Symptoms associated with elemental mercury-induced

neurotoxicity include the following: tremors, initially affecting the hands and sometimes spreading to

other parts of the body; emotional lability, often referred to as "erethism" and characterized by

irritability, excessive shyness, confidence loss, and nervousness; insomnia; neuromuscular changes (e.g.,

weakness, muscle atrophy, muscle twitching); headaches; polyneuropathy (e.g., paresthesia, stockingglove sensory loss, hyperactive tendon reflexes, slowed sensory and motor nerve conduction velocities);

and memory loss and performance deficits in test of cognitive function. At higher concentrations,

adverse renal effects and pulmonary dysfunction may also be observed.

A few studies have provided suggestive evidence for potential reproductive toxicity associated

with exposure to elemental mercury. Data from two studies in rats demonstrate developmental effects of

elemental mercury exposure. These were behavioral changes associated with both in utero and perinatal

exposure.

Inorganic Mercury

There is no evidence in humans linking exposure to mercuric chloride with carcinogenic effects. 

Data in animals are limited. Focal hyperplasia and squamous cell papillomas of the forestomach as well

as thyroid follicular adenomas and carcinomas were observed in male rats gavaged with mercuric

chloride. In the same study, evidence for an increased incidence of squamous cell forestomach

papillomas in female rats and renal adenomas and carcinomas in male mice were considered equivocal. 
ES-4

All increased tumor incidences were observed in excess of the maximum tolerated dose (MTD). In this

context, the relevance of the tumors to human health evaluation has been questioned. Results from in

vitro and in vivo tests for genotoxicity have been mixed and do not provide strong supporting data for

carcinogenicity.

There are some data indicating that mercuric chloride may be a germ cell mutagen. Positive

results have been obtained for chromosomal aberrations in multiple systems, and evidence suggests that

mercuric chloride can reach female gonadal tissue.

The most sensitive general systemic adverse effect observed following exposure to inorganic

mercury is the formation of mercuric mercury-induced autoimmune glomerulonephritis. The production

and deposition of IgG antibodies to the glomerular basement membrane can be considered the first step

in the formation of this mercuric-mercury-induced autoimmune glomerulonephritis.

Several studies in animals have evaluated the potential for developmental toxicity to occur

following exposure to various inorganic salts. While the evidence suggests that developmental effects

may occur, all of the studies have significant limitations.

Methylmercury

Three human studies that examined the relationship between methylmercury and cancer

incidence were considered extremely limited because of study design inappropriate for risk assessment or

incomplete data reporting. Evidence from animal studies provides limited evidence of carcinogenicity. 

Male ICR and B6C3F1 mice exposed orally to methylmercuric chloride were observed to have an

increased incidence of renal adenomas, adenocarcinomas, and carcinomas. Renal epithelial cell

hyperplasia and tumors, however, were observed only in the presence of profound nephrotoxicity

suggesting that the tumors may be a consequence of reparative changes to the damaged kidneys. Tumors

were observed at a single site, in a single species and sex.

Methylmercury appears to be clastogenic but not a potent mutagen. Studies have also shown

evidence that methylmercury may induce mammalian germ cell chromosome aberrations. There are a

number of studies in both humans and experimental animals that show methylmercury to be a

developmental toxicant. Neurotoxicity in offspring is the most commonly observed effect and the effect

seen at lowest exposures.

A significant body of human studies exists for evaluating the potential systemic toxicity of

methylmercury. This data base is the result of studying two large scale poisoning episodes in Japan and

Iraq as well as several epidemiological studies assessing populations that consume significant quantities

of fish. In addition, much research on the toxicity of methylmercury has been conducted in animals

including non-human primates.

The critical target for methylmercury toxicity is the nervous system. The developing fetus may

be at particular risk from methylmercury exposure. Offspring born of women exposed to methylmercury

during pregnancy have exhibited a variety of developmental neurological abnormalities, including the

following: delayed onset of walking, delayed onset of talking, cerebral palsy, altered muscle tone and

deep tendon reflexes, and reduced neurological test scores. Maternal toxicity may or may not have been

present during pregnancy for those offspring exhibiting adverse effects. For the general population, the

critical effects observed following methylmercury exposure are multiple central nervous system effects

including ataxia and paresthesia.
 The oral RfD and the inhalation RfC are estimates (with uncertainty spanning perhaps an order of 1

magnitude) of a daily exposure to the human population (including sensitive subpopulations) that is

likely to be without an appreciable risk of deleterious health effects during a lifetime.

ES-5

 A latent or silent period has been observed in some epidemiological and animal studies

indicating a delay in the onset of adverse effects. It is hypothesized this delay may be related to

methylmercury metabolism.

Sensitive Subpopulations

A susceptible population is a group that may experience more severe adverse effects at

comparable exposure levels or adverse effects at lower exposure levels than the general population. The

greater response of these sensitive subpopulations may be a result of a variety of intrinsic or extrinsic

factors. For mercury, the most sensitive subpopulations may be developing organisms. Data are also

available indicating that other factors may be associated with the identification of sensitive

subpopulations including the following: age; gender; dietary insufficiencies of zinc, glutathione, or

antioxidants; predisposition for autoimmune glomerulonephritis; and predisposition for acrodynia. More

information on sensitive subpopulations is presented in Chapter 4.

Interactions

There are data demonstrating that a number of substances affect the pharmacokinetics and/or

toxicity of mercury compounds. Of most interest is the potential interaction of selenium and mercury. 

Selenium is known to bioaccumulate in fish, so exposure to methylmercury from fish consumption may

be associated with exposure to increased levels of selenium. There are data indicating that selenium coadministered with methylmercury can form selenium-methylmercury complexes. The formation of these

complexes may temporarily prevent methylmercury-induced tissue damage but also may delay excretion

of the methylmercury. Thus, formation of selenium-methylmercury complexes may not reduce

methylmercury toxicity but rather may delay onset of symptoms. More information is needed to

understand the possible interaction of selenium with methylmercury.

There is potential for interaction between various forms of mercury and ethanol, thiol

compounds, tellurium, potassium dichromate, zinc, atrazine, and vitamins C and E.

Hazard Identification/Dose-Response Assessment

The available toxicological and epidemiological evidence was evaluated, and U.S. EPA risk

assessment guidelines and methodologies were applied to hazard identification for various endpoints;

namely, carcinogenicity, germ cell mutagenicity, developmental toxicity, and general systemic toxicity. 

Data supported quantitative assessments of systemic toxicity. For elemental mercury, an inhalation

reference concentration (RfC ) was calculated; oral reference doses (RfD ) were calculated for inorganic 1 1

mercury and methylmercury. Data for carcinogenicity of inorganic and methylmercury were judged to be

inadequate in humans and limited from animal bioassays. The carcinogenicity data for all forms of

mercury evaluated were not sufficient to support a quantitative assessment. No quantitative estimates

were done for developmental toxicity. Table ES-1 summarizes the hazard identification and doseresponse information for elemental mercury, inorganic mercury, and organic mercury. The bases for

these decisions and the methodologies applied are presented in Chapter 6.
ES-6

Table ES-1

Summary of U.S. EPA Hazard Identification/Dose-response Assessment

for Mercury and Mercury Compounds

Form Oral RfD Inhalation Cancer Cancer Germ Cell Developmental

of (mg/kg-day) RfC Weight-of- Slope Mutagenicity Toxicity

Mercury (mg/m ) evidence Factor Data Base 3

Rating Characterization

Elemental n/a 0.0003 D, not classifiable n/a Low weight of Insufficient human a b

as to human evidence evidence; sufficient

carcinogenicity animal evidence

Inorganic 0.0003 Not C, possible n/a Moderate weight Insufficient c

(mercuric verifiable human carcinogen of evidence evidence

chloride)

d

Organic 0.0001 n/a C, possible n/a High weight of Sufficient human e

(methyl- human carcinogen evidence and animal data

mercury)

 Not available; data do not support development of a value at this time. a

 Critical effect is neurological toxicity (hand tremor; increases in memory disturbances; slight subjective and objective b

 evidence of autoimmune dysfunction) in adults.

 Critical effect is renal toxicity resulting from an autoimmune disease caused by the accumulation of a hapten-mercury c

 complex in the glomerular region of the kidneys.

 Data were judged insufficient for calculation of RfC. d

 Critical effect is neurological toxicity in progeny of exposed women, RfD calculated using a benchmark dose (10%). e

Ongoing Research

While much data has been collected on the potential toxicity of mercury and mercury

compounds, much is still unknown. Two ongoing epidemiological studies are now providing critical

information on the developmental toxicity of methylmercury. One study, being conducted in the

Seychelles Islands, is evaluating dose-response relationships in a human population with dietary

exposures (fish) at levels believed to be in the range of the threshold for developmental toxicity. The

second study, conducted in the Faroe Islands, is assessing mercury exposure in a population that

consumes a relatively large quantity of marine fish and marine mammals. Children exposed to

methylmercury in utero and followed through 6 years of age have been assessed for mercury exposure

and neurological developmental. Published data from these studies are summarized in Chapter 3. 

Implications of ongoing research is discussed along with uncertainties in risk assessments in Chapter 6.

Research Needs

Specifically, information is needed to reduce the uncertainties associated with the current oral

RfDs and inhalation RfCs. More work with respect to both dose and duration of exposure would also

allow for potentially assessing effects above the RfD/RfC. Limited evidence suggests that

methylmercury and mercuric chloride are possible human carcinogens. Data are not sufficient to classify

the potential carcinogenicity of elemental mercury. Research on mode of action in induction of tumors at

high mercury dose will be of particular use in defining the nature of the dose response relationship for

carcinogenicity. At this time data have been judged insufficient for calculation of quantitative

developmental toxicity estimates for elemental and inorganic mercury; research toward this end should

be encouraged. While some pharmacokinetic models have been developed additional work to ensure the
ES-7

applicability of these to risk assessment should be pursued. In particular work aimed at validation of a

fetal pharmacokinetic model and research in support of toxicokinetics will be useful.

Conclusions

The following conclusions progress from those with greater certainty to those with lesser

certainty.

 The three forms of mercury discussed in this Report can present a human health hazard.

 Neurotoxicity is the most sensitive indicator of adverse effects in humans exposed to

elemental mercury and methylmercury.

 Immune-mediated kidney toxicity is the most sensitive indicator of toxic effects of

exposure to inorganic mercury. This judgement is largely based on results in

experimental animals.

 Methylmercury is a developmental toxicant in humans.

 Methylmercury is likely to be a human germ cell mutagen. This judgement is based on

data from human studies, genetic toxicology studies in animals and a consideration of the

pharmacokinetics of methylmercury.

 An RfD for ingested methylmercury based on neurotoxic effects observed in Iraqi

children exposed in utero is 1 x 10 mg/kg-day. The threshold estimate derived using a -4

benchmark dose approach is not model dependent (polynomial vs. Weibull). The

estimate is not much affected by data grouping, but is dependent on response

classification and on parameters used in determination of ingestion relative to measured

mercury in hair.

 An RfC for inhaled elemental mercury based on neurotoxic effects in exposed workers is

3 x 10 mg/m . -4 3

 An RfD for ingested inorganic mercury based on immune-mediated kidney effects in

Brown-Norway rats is 3 x 10 mg/kg-day. -4

 Elemental mercury is a developmental toxicant in experimental animals. If the

mechanisms of action producing developmental toxicity in animals occur in humans,

elemental mercury is very likely to produce developmental effects in exposed human

populations. U.S. EPA has made no estimate of dose response for developmental effects

of elemental mercury.

 Methylmercury and inorganic mercury produce tumors in experimental animals at toxic

doses. If the mechanisms of action which induced tumors in the animal models could

occur in humans, it is possible that tumors could be induced in exposed humans by these

forms of mercury. It is likely, however, that cancer would be induced only after

mercury exposures in excess of those producing other types of toxic response.
ES-8

There are many uncertainties associated with this analysis, due to an incomplete understanding

of the toxicity of mercury and mercury compounds. The sources of uncertainty include the

following:

 The data serving as the basis for the methylmercury RfD were from a population

ingesting contaminated seed grain. The nutritional status of this group may not be

similar to that of U.S. populations. The exposure was for a short albeit critical period of

time. It is likely that there is a range of response among individuals to methylmercury

exposure. The selenium status of the exposed Iraqi population is not certain, nor is it

established the extent to which selenium has an effect on mercury toxicity.

 There was no NOAEL (no-observable-adverse-effect level) for estimation of a threshold

for all developmental endpoints. A benchmark was estimated using a Weibull model on

grouped data. Use of an estimate other than the 95% lower limit on 10% response

provides alternate estimates. Other modeling approaches using data which have not been

grouped provide similar estimates. Benchmark doses, NOAELs, LOAELs, from other

human studies provide support for the benchmark used in the RfD.

 Ingestion levels of methylmercury associated with measured mercury in hair were

estimated based on pharmacokinetic parameters derived from evaluation of the extant

literature. Use of other plausible values for these parameters results in (relatively small)

changes in the exposure estimate.

 While there are data to show that the developing fetus is more susceptible to

methylmercury toxicity than adults, there are not sufficient data to support calculation of

a separate RfD for children (vs. adults).

 The RfD for inorganic mercury is based on data in experimental animals; there is

uncertainty in extrapolation to humans. It is thought that these animals constitute a good

surrogate for a sensitive human subpopulation. The data were from less than lifetime

exposures; there is uncertainty in extrapolation to a lifetime RfD. There was no NOAEL

in the studies; there is uncertainty in extrapolation to a NOAEL or in estimation of a

threshold for effects in animals.

 The RfC for elemental mercury was based on studies in exposed workers for which there

is no reported NOAEL; there is uncertainty in estimating the no effect level in these

populations. There is uncertainty as to whether reproductive effects could be occurring

at lower exposure levels than those which produced the observed neurotoxicity.

 There are insufficient data to determine whether elemental mercury induces carcinogenic

effects in experimental animals.

 Data are not sufficient to judge if elemental and inorganic mercury are germ cell

mutagens.

 U.S. EPA did not formally evaluate data on mercury for reproductive effects.
ES-9

To improve the risk assessment for mercury and mercury compounds, U.S. EPA would need the

following:

 Results from ongoing studies in human populations with measurable exposure to

methylmercury.

 Results for immune-mediated kidney effects from lifetime studies of sensitive animals

exposed to inorganic mercury. Definitive data from human studies on effects of

exposure to inorganic mercury.

 Data on inhalation effects of inorganic mercury exposure.

 Dose response data for developmental effects of elemental and inorganic mercury.

 Reproductive studies and analysis for all forms of mercury.

 Data on mode of action of inorganic and methylmercury tumor induction.

 Validated physiologically-based pharmacokinetic models for mercury which include a

fetal component.

Based on the extant data and knowledge of developing studies, the following outcomes can be

expected:

 Human populations exposed to sufficiently high levels of elemental mercury will have

increased incidence of neurotoxic effects.

 Human populations exposed to sufficiently high levels of methylmercury either in utero

or post partum will have increased incidence of neurotoxic effects.

 Human populations exposed to sufficiently high levels of inorganic mercury will have

increased incidence of systemic effects including immune-mediated kidney effects.

 The RfDs and RfC calculated by U.S. EPA for systemic toxic effects of mercury are

expected to be amounts of exposure that can be incurred on a daily basis for a lifetime

without anticipation of adverse effects. This expectation is for populations including

susceptible subpopulations.

 The RfDs are protective against carcinogenic effects; tumor induction in animals was

observed only at doses likely to produce systemic toxic effects.
1-1

1. INTRODUCTION

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

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

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

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

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

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

controls.

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

Congress. The eight volumes are as follows:

I. Executive Summary

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

III. Fate and Transport of Mercury

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

V. Health Effects of Mercury and Mercury Compounds

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

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

United States

VIII. An Evaluation of Mercury Control Technologies and Costs

This volume (Volume V) addresses the potential human health effects associated with exposure

to mercury. It summarizes the available human and animal studies and other supporting information

relevant to the toxicity of mercury and mercury compounds in humans. It also summarizes U.S. EPA's

current overall assessments of hazard and quantitative dose-response for various categories of toxic

effects. This volume presents data relevant to assessment of potential effects on human health for

elemental mercury, inorganic mercury and methylmercury. Organic mercury compounds other than

methylmercury are generally not considered in this volume. Chapter 2 discusses the toxicokinetics of

mercury, including information on absorption, distribution, metabolism and excretion. Chapter 3 is a

summary of the toxicity literature for mercury. It is organized into three main subsections, corresponding

to elemental mercury, inorganic mercury and methylmercury. Within each of these subsections, the

study data are presented according to the effect type (e.g., death, renal toxicity, developmental toxicity,

cancer). For each effect type, separate summary tables in similar formats are used to present the

available data from human epidemiological studies, human case studies, and animal studies.

Chapter 6, Hazard Identification and Dose-Response Assessment, presents U.S. EPA's

assessments of the hazard presented by various forms of mercury and, where possible, the quantitative

dose-response information that is used in risk assessments of mercury. Chapters 4 and 5 briefly discuss

populations with increased susceptibility to mercury and interactions between exposure to mercury and

other substances. Ongoing research and research needs are described in Chapter 7, and Chapter 8 lists

the references cited. Appendix A documents the dose conversion equations and factors used. Appendix

B consists of RfD, RfC and cancer risk summaries for U.S. EPA’s Integrated Risk Information System

(IRIS). Appendix C lists the participants of a U.S. EPA-sponsored workshop on mercury issues held in

1987. Appendix D presents an analysis of uncertainty and variability in the methylmercury human

effects threshold estimate.
2-1

2. TOXICOKINETICS

This chapter describes the toxicokinetics (i.e., absorption, distribution, metabolism and

excretion) of mercury and mercury compounds in the body. Biomarkers of exposure and methods of

analysis for measuring mercury levels in biological samples are discussed. The biotransformation of

mercury in the environment is discussed in Volume III.

The absorption of elemental mercury vapor occurs rapidly through the lungs, but it is poorly

absorbed from the gastrointestinal tract. Oral absorption of inorganic mercury involves absorption

through the gastrointestinal tract; absorption information for the inhalation route is limited. 

Methylmercury is rapidly and extensively absorbed through the gastrointestinal tract.

Once absorbed, elemental mercury is readily distributed throughout the body; it crosses both

placental and blood-brain barriers. Elemental mercury is oxidized to inorganic divalent mercury by the

hydrogen peroxidase-catalase pathway, which is present in most tissues. The oxidation of elemental

mercury to the inorganic mercuric cation in the brain can result in retention in the brain. Inorganic

mercury has poor lipophilicity and a reduced capacity for penetrating the blood-brain or placental

barriers. Once elemental mercury crosses the placental or blood-brain barriers and is oxidized to the

mercuric ion, return to the general circulation is impeded, and mercury can retained in brain tissue. 

Recent studies indicate that transport and distribution of methylmercury is carrier-mediated. 

Methylmercury penetrates the blood-brain and placental barriers, can be converted to mercuric ion, and

may accumulate in the brain and fetus.

The elimination of elemental mercury occurs via the urine, feces and expired air. Exposure to

mercuric mercury results in the elimination of mercury in the urine and feces. Methylmercury is excreted

primarily in the feces (mostly in the inorganic form) by humans.

2.1 Absorption

2.1.1 Elemental Mercury

2.1.1.1 Inhalation

Elemental mercury vapors are readily absorbed through the lungs. Studies in human volunteers

have shown that approximately 75–85% of an inhaled dose of elemental mercury vapor was absorbed by

the body (Nielsen-Kudsk 1965; Oikawa et al. 1982; Teisinger and Fiserova-Bergerova 1965; Hursh 1985;

Hursh et al. 1985). The high lipid solubility of elemental mercury vapor relative to its vapor pressure

favors its rapid diffusion across alveolar membranes and dissolution in blood lipids (Berlin et al. 1969b).

2.1.1.2 Oral

Liquid metallic mercury is very poorly absorbed from the gastrointestinal tract. In rats, less than

0.01% of an ingested dose of metallic mercury was absorbed (Bornmann et al. 1970). The release of

mercury vapor from liquid elemental mercury in the gastrointestinal tract and the subsequent absorption

of the released vapor is limited by reaction of the mercury with sulfur to form mercuric sulfide. The

mercuric sulfide coats ingested metallic mercury, preventing release of elemental vapor (Berlin 1986).
2-2

2.1.1.3 Dermal

Elemental mercury vapor is absorbed through the skin of humans at an average rate of 0.024 ng

Hg/cm (skin) for every one mg/m in the air (Hursh et al. 1989). This rate of dermal absorption is 2 3

sufficient to account for less than 3% of the total amount absorbed during exposure to mercury vapor

(greater than 97% of the absorption occurs through the lungs). Dermal absorption of liquid metallic

mercury has been demonstrated in experimental animals (Schamberg et al. 1918); however, the extent of

absorption was not quantified. Koizumi et al. (1994) measured mercury absorption through the skin of

F344 rats exposed to solutions of industrially generated dust containing mercury. After 3 days mean

blood concentration in dust-exposed rats was 15.5 g/L compared to 3 g/L for saline controls.

2.1.2 Inorganic Mercury

2.1.2.1 Inhalation

There is limited information suggesting that absorption occurs after inhalation of aerosols of

mercuric chloride. Clarkson (1989) reported absorption to be 40% in dogs via inhalation. Inhalation

exposure of rats to an aerosol of a 1% mercuric chloride solution for 1 hour/day, 4 days/week for 2

months resulted in retention of 5–6 g HgCl/HR aerosol/100 g body weight or approximately 37-44 μg

Hg/kg-day (Bernaudin et al. 1981). The authors prepared the aerosol “with reference to the maximum

allowable air concentrations (0.10 mg Hg/m ) for a man”. Retention was defined at the end of exposure 3

as total mercury in the rat carcass minus skin and hair. It is unknown to what extent the amount retained

represented absorption through the lungs or absorption of material cleared from the respiratory tract by

mucociliary activity and ultimately swallowed.

2.1.2.2 Oral

The absorption of mercuric mercury from the gastrointestinal tract has been estimated at

approximately 7–15% in human volunteers following oral administration of radiolabeled inorganic

mercury (Miettinen 1973; Rahola et al. 1973). Recent data from studies in mice, however, suggest that

"true" absorption may be closer to 20% but appears lower due to intestinal pH, compound dissociation,

age, diet, rapid biliary secretion and excretion in the feces (Kostial et al. 1978; Nielsen 1992). Because

the excretion of absorbed mercury is rapid, mercury levels detected in the gastrointestinal tract most

likely represent both unabsorbed and excreted mercury in the studies by Miettinen (1973) and Rahola et

al. (1973). The absorption of mercuric chloride from the gastrointestinal tract is not believed to depend

on any specific transport mechanism, reactive sulfhydryl groups, or oxidative metabolism (Foulkes and

Bergman 1993). Rather, uptake appears to result from an electrostatic interaction with the brush border

membrane and limited passive diffusion. Several factors have been identified that modulate absorption

of mercuric mercury from the gastrointestinal tract. At high doses, the corrosive action of mercuric

chloride may increase its uptake by breaking down membrane barriers between the ions and the blood. 

Increases in intestinal pH also increase absorption (Endo et al. 1990). Increased uptake also occurs in

neonates (Kostial et al. 1978). The increased absorption in neonates is believed to be due in part to the

milk diet of neonates (increased absorption was observed in adults given a milk diet) and in part to the

increased pinocytotic activity in the gastrointestinal tract that occurs in the very young (Kostial et al.

1978). Diffusion through aqueous channels present in the immature brush border of neonates has also

been suggested to account for the greater absorption in the very young (Foulkes and Bergman 1993).

Absorption of mercuric salts from the gastrointestinal tract varies with the particular salt

involved. Absorption decreases with decreasing solubility (Endo et al. 1990). For example, the poorly
2-3

soluble salt mercuric sulfide is not absorbed from the gastrointestinal tract as well as the more soluble

mercuric chloride salt (Sin et al. 1983).

Mercurous salts in the form of calomel (long in use as a therapeutic agent) are insoluble in water

and are poorly absorbed from the gastrointestinal tract (Clarkson 1993a). Long term use of calomel,

however, has resulted in toxicity in humans (Davis et al. 1974).

2.1.2.3 Dermal

Dermal absorption of mercuric chloride has been observed in treated guinea pigs (Skog and

Wahlberg 1964). Approximately 2–3% of an applied dose was absorbed during a 5-hour period. 

Absorption was measured both by disappearance of the applied compound and by appearance in kidney,

liver, urine and blood.

2.1.3 Methylmercury

2.1.3.1 Inhalation

Inhaled methylmercury vapors are absorbed through the lungs. Fang (1980) did not measure

percent absorbed but showed a correlation between tissue mercury levels and both exposure level and

duration in rats exposed to radioactively labelled methylmercury vapor.

2.1.3.2 Oral

Methylmercury is efficiently absorbed from the gastrointestinal tract. Approximately 95% of

methylmercury in fish ingested by volunteers was absorbed from the gastrointestinal tract (Aberg et al.

1969; Miettinen 1973). Similarly, when radiolabeled methylmercuric nitrate was administered in water

to volunteers, uptake was greater than 95% (Aberg et al. 1969).

Reports of the percentage of absorbed methylmercury distributed to the blood range from 1% to

10%. Following the ingestion of a single meal of methylmercury-contaminated fish, Kershaw et al.

(1980) found that blood accounted for 5.9% of absorbed methylmercury, while Miettinen et al. (1971)

found an initial value of 10%, decreasing to about 5% over the first 100 days. In a population that

chronically ingested fish with high methylmercury levels, approximately 1% of the absorbed dose was

distributed to the blood (Sherlock et al. 1982).

2.1.3.3 Dermal

Dermal absorption of the methylmercuric cation (CH H3 g) (as the dicyandiamide salt) has also +

been observed in treated guinea pigs (Skog and Wahlberg 1964). Approximately 3–5% of the applied

dose was absorbed during a 5-hour period. Absorption was measured both by disappearance of the

applied compound and by appearance in kidney, liver, urine and blood.
2-4

2.2 Distribution

2.2.1 Elemental Mercury

Because of its lipophilicity, absorbed elemental mercury vapor readily distributes throughout the

body, crossing the blood-brain barrier in humans (Hursh et al., 1976; Nordberg and Serenius, 1969) and

the placenta in rats and mice (Clarkson et al., 1972). The distribution of absorbed elemental mercury is

limited primarily by the oxidation of elemental mercury to the mercuric ion and reduced ability of the

mercuric ion to cross membrane barriers. The oxidation is sufficiently slow, however, to allow

distribution to all tissues and organs. Once it is oxidized to the mercuric ion, it is indistinguishable from

Hg from inorganic sources (i.e., the highest levels of mercury accumulate in the kidneys) (Hursh et al. 2+

1980; Rothstein and Hayes 1964). Based on an in vitro study by Hursh et al. (1988), oxidation of

mercury in the blood is slow and, therefore, inhaled mercury reaches the brain primarily unoxidized (i.e.,

as dissolved vapor) and is available for rapid penetration into brain cells. Once in the brain, oxidation of

elemental mercury to mercuric mercury in the brain enhances for the accumulation of mercury in these

tissues (Hursh et al. 1988; Takahata et al. 1970). For example, ten years after termination of exposure,

miners exposed to elemental mercury vapor had high concentrations of mercury (120 ppm) in the brain

(Takahata et al. 1970). A similar effect occurs when elemental mercury reaches the fetus and (after

oxidation) accumulates in the tissues as inorganic mercury (Dencker et al. 1983).

In the blood, elemental mercury initially distributes predominantly to the red blood cells; at

20 minutes, 98% of the mercury in the blood is found in the red blood cells. Several hours following

parenteral, oral or inhalation exposure, however, a stable ratio of red blood cell mercury to plasma

mercury of approximately 1:1 is established (Gerstner and Huff, 1977; Clarkson, 1972; Cherian et al.,

1978). The rise in plasma mercury levels was suggested to be due to binding to protein sulfhydryl groups

by mercuric mercury formed when the elemental mercury was oxidized.

2.2.2 Inorganic Mercury

In contrast to elemental mercury vapor and methylmercury, mercuric mercury does not penetrate

the blood-brain or placental barriers easily. Levels of mercury observed in the rat brain after injection of

mercuric nitrate were 10-fold lower than after inhalation of an equivalent dose of elemental mercury

vapor (Berlin et al. 1969a). Similarly, mercuric mercury shows only limited ability to penetrate to the

fetus (Garrett et al. 1972). Mercuric mercury does, however, accumulate in the mouse placenta (Berg

and Smith 1983; Mitani et al. 1978; Suzuki et al. 1984). In the blood, the mercuric ion is bound to

sulfhydryl groups present in the plasma and erythrocytes. The ratio of human red blood cell mercuric

mercury to plasma mercuric mercury is approximately 1:1 (0.53:1.20) (Hall et al. 1994). The half-life in

blood for humans was reported to range from 19.7 to 65.6 days in a study of five subjects treated with i.v.

mercuric nitrate (Hall et al. 1994). From the blood, mercuric mercury initially distributes to liver, but the

highest levels are generally observed in the kidneys (Newton and Fry 1978). With time after exposure,

accumulation in the kidneys may account for up to 90% of the total body burden (Rothstein and Hayes

1960). The mercury levels in the kidney are dose dependent, with increasing amounts occurring with

higher administered dose levels (Cember, 1962). The highest concentration of mercuric mercury in the

kidneys is found in the proximal tubules. Mercuric mercury induces metallothionein production in the

kidneys (Piotrowski et al. 1974). The high metallothionein levels in the kidneys may contribute to the

kidney's accumulation of mercuric mercury (Piotrowski et al. 1973).

In neonates, lower proportions of mercuric mercury distribute to the kidneys than in adult

animals (Jugo 1976). This results in higher distribution to other tissues. The protective blood-brain
2-5

barrier is incomplete in fetal and neonatal animals, which may also contribute to the increased mercury

levels in immature brain. For example, the higher levels in the neonatal brain of rats and guinea pigs are

believed to be associated with the decrease in renal sequestration of the mercuric ion (Jugo 1976;

Yoshida et al. 1989). The higher levels observed in the livers of rat neonates may be attributable to

increased distribution to organs other than the kidney as well as to higher levels of neonatal hepatic

metallothionein (Daston et al. 1986). 

2.2.3 Methylmercury

Methylmercury is distributed throughout the body, easily penetrating the blood-brain and

placental barriers in humans and animals (Clarkson 1972; Hansen 1988; Hansen et al. 1989; Nielsen and

Andersen 1992; Soria et al. 1992; Suzuki et al. 1984). By contrast with elemental mercury, studies in rats

indicate that methylmercury transport into tissues is mediated by the formation of a methylmercurycysteine complex (Aschner and Aschner 1990; Tanaka et al. 1991, 1992; Kerper et al. 1992). The

complex is structurally similar to methionine and is transported into cells via a widely distributed neutral

amino acid carrier protein. Methylmercury associates with water-soluble molecules (e.g., proteins) or

thiol-containing amino acids because of the high affinity of the methylmercuric cation (CH H3 g) for the +

sulfhydryl groups (SH) . Complexes of methylmercury with cysteine have been identified in blood, liver -

and bile of rats (Aschner and Aschner 1990).

Al-Shahristani and Shihab (1974) calculated a “biological half-life” of methylmercury in a study

of 48 male and female subjects who had ingested seed grain contaminated by organic mercurials. The

half-life ranged from 35 to 189 days with a mean of 72 days; it was determined from distribution of

mercury along head hair.

The blood half-life is 49–164 days in humans (Aberg et al. 1969; Miettinen et al. 1971) and

10–15 days in monkeys (Rice et al. 1989). Smith et al. (1994) determined a blood half-life of 32–60 days

in a study of seven adult males given i.v. methylmercury. In the blood, methylmercury is found

predominantly in the red blood cells (Kershaw et al. 1980; Thomas et al. 1986). In humans, the ratio of

red blood cell methylmercury to plasma methylmercury is approximately 20:1. This ratio varies in

animal species; the ratio is approximately 20:1 in primates and guinea pigs, 7:1 in mice, greater than

100:1 in rats and 42:1 in cats (Hollins et al. 1975; Magos 1987).

The clinical significance of the differences in the distribution of various forms of mercury in the

blood is that it permits diagnosis of the type of mercury to which an individual has been exposed. Shortchain alkyl mercury compounds such as methylmercury or ethyl mercury are very stable in the body,

whereas long-chain compounds may be metabolized over time to the mercuric ion. The mercury

distribution in the blood, therefore, may shift from a distribution characteristic of methylmercury to one

more suggestive of inorganic mercury (Berlin 1986; Gerstner and Huff 1977).

Mercury has been found in the umbilical cord of human newborns at levels comparable to

maternal blood levels (Grandjean et al. 1992a). For lactating mothers, the clearance of mercury from the

blood appears to be faster than for non-lactating women. Lactating individuals have a blood half-life of

42 days compared to 75 days for non-lactating females among a group of people who had consumed

contaminated seed grain (Greenwood et al. 1978). This finding may be due to excretion of mercury via

the milk, increased food intake by mothers (which enhances biliary excretion) and/or altered hormonal

patterns in lactating mothers (which affect the excretion pattern).
2-6

Methylmercury transport across the blood-brain barrier in rats may involve an amino acid carrier

(Kerper et al. 1992). Following acute exposure to methylmercury, most of the mercury in the brain is in

the organic form; however, with chronic exposures, a greater amount of the mercury in the brain is in the

inorganic form, suggesting that the rate of demethylation increases with long-term exposure (Aschner

and Aschner 1990). Rice (1989a, 1989b) demonstrated that tissue half-life in the brain may be

significantly longer than the blood half-life for methylmercury.

The bioaccumulation of methylmercury can be affected by age and sex (Thomas et al. 1982,

1986, 1988). After administration of methylmercury to rats, the females had higher peak levels of

mercury in the kidneys, primarily as methylmercury, compared to the males; inorganic mercury levels did

not differ significantly between the sexes (Thomas et al. 1986). Accumulation of mercury in the body is

also found to be higher in neonatal rats (Thomas et al. 1988) than in adult rats (Thomas et al. 1982). Ten

days after administration of methylmercury, 94% of the dose was still detected in neonates while 60%

was retained in adults (Thomas et al. 1988). The longer retention of mercury in the neonates may be

attributed to various factors including the high amount of mercury accumulated in the pelt of the

neonates due to lack of clearance (Thomas et al. 1988) and the lack of a fully developed biliary transport

system in the neonates (Ballatori and Clarkson 1982).

2.3 Metabolism

2.3.1 Elemental Mercury

Elemental mercury dissolved in the blood is rapidly oxidized in red blood cells to mercuric

mercury by catalase in the presence of hydrogen peroxide (Halbach and Clarkson 1978). Catalase is

found in many tissues, and oxidation by this pathway probably occurs throughout the body (NielsenKudsk 1973). The pathway is saturable, however, and hydrogen peroxide production is the rate-limiting

step (Magos et al. 1989). Blood and tissue levels of mercuric mercury following exposure to high

concentrations of elemental mercury are, therefore, lower than would be expected based on levels

observed following exposure to low levels.

2.3.2 Inorganic Mercury

Several investigators have observed exhalation of elemental mercury vapor after oral

administration of mercuric mercury to rats and mice, indicating that mercuric mercury in the body can be

reduced to elemental mercury (Clarkson and Rothstein 1964; Dunn et al. 1981a, 1981b; Sugata and

Clarkson 1979). The reduction of mercuric ion to elemental mercury may occur via cytochrome c,

NADPH and NADH, or a superoxide anion produced by the xanthine-xanthine oxidase system (Ogata et

al. 1987). There is no evidence that mercuric mercury is methylated to form methylmercury in

mammalian cells. The studies of Rowland et al. on the intestinal flora of the Wistar rat show that

microbes are responsible for at least a portion of mercuric chloride methylation in the gut.

Mercurous mercury is unstable in biological fluids and rapidly disassociates to one molecule of

elemental mercury and one ion of mercuric mercury (Clarkson 1972).

2.3.3 Methylmercury

Methylmercury in the body is relatively stable and is only slowly demethylated to form mercuric

mercury in rats (Norseth and Clarkson 1970). The demethylation appears to occur in tissue macrophages

(Suda and Takahashi 1986), intestinal microflora (Nakamura et al. 1977; Rowland et al. 1980) and fetal
2-7

liver (Suzuki et al. 1984). In vitro demethylation has been reported to involve hydroxyl radicals

produced by cytochrome P-450 reductase (Suda and Hirayama 1992) or hypochlorous acid scavengers

(Suda and Takahashi 1992). Organic mercury compounds with longer alkyl chains are more readily

metabolized over time to the mercuric ion (Berlin, 1986).

Methylmercury metabolism may be related to the latent or silent period observed in

epidemiological studies from two methylmercury poisonings. During the latent period, both during and

after the cessation of exposure, the patient feels no untoward effects. It is possible that a number of

biochemical changes may take place in parallel during this period, and some may not be causatively

related to the clinical outcome. Ganther (1978) has hypothesized that the carbon-mercury bond in

methylmercury undergoes homolytic cleavage to release methyl free radicals. The free radicals are

expected to initiate a chain of events involving peroxidation of lipid constituents of the neuronal cells. 

The onset of symptoms is delayed for the period of time that cellular systems are able to prevent or repair

effects of lipid peroxidation. When the cellular defense mechanisms are overwhelmed, rapid and

progressive degeneration of the tissue results. In the Iraqi poisoning incident, the latent period before

toxic signs were noted varied from a matter of weeks to months. By contrast, in the Japanese poisoning

incident, the latency was as long as a year or more. The difference in duration of the latent period may in

part be due to the presence of selenium in the fish ingested by the Japanese population. The role of

selenium in mercury toxicity is discussed further in Chapter 5.

2.4 Excretion

2.4.1 Elemental Mercury

Excretion of mercury after exposure to elemental mercury vapor may occur via exhaled air,

urine, feces, sweat and saliva. The pattern of excretion of elemental mercury changes as elemental

mercury is oxidized to mercuric mercury. During and immediately after an acute exposure, when

dissolved elemental mercury is still present in the blood, glomerular filtration of dissolved mercury vapor

occurs, and small amounts of mercury vapor can be found in the urine (Stopford et al. 1978). Mercury

vapor present in the blood may also be exhaled; human volunteers exhaled approximately 7% of the

retained dose within the first few days after exposure (Hursh et al. 1976). The half-life for excretion via

the lungs is approximately 18 hours. Approximately 80% of the mercury accumulated in the body is

eventually excreted as mercuric mercury. As the body burden of mercury is oxidized from elemental

mercury to mercuric mercury, the pattern of excretion becomes more similar to mercuric mercury

excretion. The majority of the excretion of mercuric mercury occurs in the feces and urine (Cherian et al.

1978). During the first few days after exposure of humans to mercury vapor, approximately four times

more mercury was excreted in the feces than in the urine (Cherian et al. 1978). With time, as the relative

mercury content of the kidneys increases, excretion by the urinary route also increases (Rothstein and

Hayes 1964). Tissue levels of mercury decrease at different rates, but the half-life for excretion of

whole-body mercury in humans (58 days) is estimated to be approximately equal to the half-life of

elimination from the kidneys (64 days), where most of the body burden is located (Hursh et al. 1976). 

Excretion via the urine may be increased if mercury-induced damage of the renal tubular epithelium has

happened and exfoliation of damaged mercury-containing cells occurs (Magos 1973).

Excretion via sweat and saliva are thought to contribute only minimally to total excretion under

normal circumstances. In workers who have perspired profusely, however, the total amount of mercury

excreted in the sweat during 90 minutes ranged from 50% to 200% of that found in a 16-hour composite

sample of urine (Lovejoy et al. 1974).
2-8

2.4.2 Inorganic Mercury

Because of the poor absorption of orally administered mercuric mercury, the majority (85%) of

an ingested dose in humans is excreted in the feces within a few days after administration (Miettinen

1973). Hall et al. (1994) showed that for five adult male volunteers given i.v. mercuric nitrate and

evaluated for 70 days, 6.3-35% of the dose was excreted in urine and 17.9-38.1% in feces. For absorbed

inorganic mercury, the half-life for excretion has been estimated to be 40 days (Rahola et al. 1973) and

67 days with a range of 49-96 days (Hall et al. 1994). Information on the routes of excretion for

absorbed inorganic mercury are limited, but excretion would be expected to be similar to that of

inorganic mercury formed in rats by the oxidation of elemental mercury (Rothstein and Hayes 1964). 

The majority of absorbed inorganic mercury is excreted in the urine (Berlin 1986).

Glomerular filtration is not thought to contribute substantially to urinary excretion of mercuric

mercury (Cherian et al. 1978). Rather, mercuric mercury is excreted in the urine primarily as sulfhydryl

conjugates (with cysteine or N-acetylcysteine) actively transported into the tubular lumen. Urinary levels

correlate with renal mercury concentrations rather than blood mercury levels.

Fecal excretion of mercury occurs as the result of excretion in the saliva, secretion through the

epithelium of the small intestines and colon and secretion in the bile (Berlin 1986). Secretion of

mercuric mercury in the bile is believed to result from active transport of a mercury-glutathione complex

across the canalicular membrane via the glutathione carrier (Ballatori and Clarkson 1982).

Mercuric mercury may also be excreted in breast milk during lactation (Yoshida et al. 1992). 

The levels in breast milk are proportional to the plasma content. In maternal guinea pigs, milk levels

were approximately half of that found in plasma. After termination of exposure, however, mercury levels

in milk decreased at a slower rate than plasma mercury levels.

2.4.3 Methylmercury

Like inorganic mercury, methylmercury has a relatively long half-life of approximately 70–80

days in the human body (Aberg et al. 1969; Bernard and Purdue 1984; Miettinen 1973). Recently a

shorter half-life of 44 days was estimated by Smith et al. (1994) in their study of seven adult males

treated i.v. with methylmercury. In this study methylmercury and inorganic mercury concentrations in

blood and excreta were determined separately based on differential extractability into benzene. The

predominant species in the blood was methylmercury; there was no detectable methylmercury in the

urine.

The long half-life of methylmercury in the body is due, in part, to reabsorption of methylmercury

secreted into the bile (hepato-biliary cycling) (Norseth and Clarkson, 1971). In this cycle,

methylmercury forms a complex with glutathione in the hepatocyte, and the complex is secreted into the

bile via a glutathione carrier protein (Clarkson, 1993b). The methylmercury-glutathione complex in the

bile may be reabsorbed from the gallbladder and intestines into the blood. When microorganisms found

in the intestines demethylate methylmercury to form mercuric mercury, this cycle is broken, and fecal

excretion of mercury from methylmercury occurs (Rowland et al. 1980). Mercuric mercury is poorly

absorbed from the intestines, and that which is not reabsorbed is excreted in the feces. In humans,

approximately 90% of the absorbed dose of methylmercury is excreted in the feces as mercuric mercury. 

Excretion via the urine is minor but slowly increases with time; at 100 days after dosing, urinary

excretion of mercury accounted for 20% of the daily amount excreted. The urinary excretion of mercury

may reflect the deposition of demethylated mercury in the kidneys and its subsequent excretion.
2-9

In animals, the predominant route of methylmercury elimination also is the feces (Farris et al.

1993; Hollins et al. 1975; Thomas et al. 1987). As in humans, biliary excretion of methylmercury and its

demethylation in gastrointestinal flora have been reported in rats (Farris et al., 1993). After a single oral

dose of methylmercury, the major elimination route was the feces (65% of the administered dose as

inorganic mercury and 15% of the administered dose as methylmercury) and the minor route was urine

(1% of the administered dose as inorganic mercury and 4% of the administered dose as methylmercury)

(Farris et al. 1993).

In rat and monkey neonates, excretion of methylmercury is severely limited (Lok 1983; Thomas

et al. 1982). In rats dosed prior to 17 days of age, essentially no mercury was excreted (Thomas et al.

1982). By the time of weaning, the rate of excretion had increased to adult levels. The failure of

neonates to excrete methylmercury may be associated with the inability of suckling infants to secrete bile

(Ballatori and Clarkson 1982) and the decreased ability of intestinal microflora to demethylate

methylmercury during suckling (Rowland et al. 1977).

Methylmercury is also excreted in breast milk (Bakir et al. 1973; Sundberg and Oskarsson 1992). 

The ratio of mercury in breast milk to mercury in whole blood was approximately 1:20 in women

exposed to methylmercury via contaminated grain in Iraq between 1971 and 1972 (Bakir et al. 1973). 

Evidence from the Iraqi poisoning incident also showed that lactation decreased blood mercury clearance

half-times from 75 days in males and nonlactating females to 42 days in lactating females; the faster

clearance due to lactation was confirmed in mice (Greenwood et al. 1978). In mice, of the total mercury

in the breast milk, approximately 60% was estimated to be methylmercury. Skerfving (1988) has found

that 16% of mercury in human breast milk is methylmercury. Studies in animals indicate that the

mercury content of breast milk is proportional to the mercury content of plasma (Sundberg and

Oskarsson, 1992; Skerfving, 1988).

2.5 Biological Monitoring

This section describes the various biological media most frequently used when assessing mercury

exposure. In addition, this section describes the available analytical methods for measuring mercury in

biological samples. Reference values for mercury in standard biological media from the general

population are shown in Table 2-1. These values represent total mercury, not individual mercury species. 

For hair and blood, these have been indexed to fish consumption as the most common route of exposure

in humans.

2.5.1 Elemental Mercury

Blood and urinary mercury are common to assess occupational mercury exposure.

2.5.1.1 Blood

In workers chronically exposed to mercury vapor, a good correlation was observed between

intensity of mercury vapor exposure and levels of mercury in the blood at the end of a workshift (Roels et

al. 1987). The usefulness of blood as a biomarker for exposure to elemental mercury depends on the time

elapsed since exposure and the level of exposure. For recent, high-level exposures, whole blood analysis

may be used to assess exposure (Clarkson et al. 1988). Mercury in the blood peaks rapidly, however, and

decreases with an initial half-life of approximately two to four days (Cherian et al. 1978). Thus,

evaluation of blood mercury is of limited value if a substantial amount of time has elapsed since

exposure. Also, dietary methylmercury contributes to the amount of mercury measured in blood. At low
2-10

levels of elemental mercury exposure, the contribution of dietary methylmercury to the total blood

mercury may be high relative to that of the inhaled mercury, limiting the sensitivity of this biomarker. 

Several studies have 

Table 2-1

Reference Values for Total Mercury Concentrations in Biological Media for

the General Population

Medium Mercury Concentration Reference

Whole blood 1–8 μg/L WHO (1990)

 Fish consumption: Brune (1991)

 No fish meals 2.0 μg/L

 2 meals/week 4.8 μg/L

 2-4 meals/week 8.4 μg/L

 more than 4 meals/week 44.4 μg/L

 2 μg/L Nordberg et al. (1992)

Urine 4–5 μg/L WHO (1990)

Scalp hair 2 μg/g WHO (1990)

 Fish consumption: Airey (1983)

 once/mo 1.4 μg/g

 once/2 wk 1.9 μg/g

 once/wk 2.5 μg/g

 once/day 11.6 μg/g

separated whole blood into its plasma and erythrocyte fractions in order to evaluate potential

confounding factors due to the presence of methylmercury (95% of methylmercury is found in the red

blood cell). Some published values indexed to fish consumption are in Table 2-1.

2.5.1.2 Urine

Urinary mercury is thought to indicate most closely the mercury levels present in the kidneys

(Clarkson et al. 1988). For most occupational exposures, urinary mercury has been used to estimate

exposure. In contrast to blood mercury levels, urinary mercury peaks approximately 2–3 weeks after

exposure and decreases at a much slower rate with a half-life of 40–60 days for short-term exposures and

90 days for long-term exposures (Barregard et al. 1992; Roels et al. 1991). The urine remains, therefore,

a more appropriate indicator for longer exposures than blood samples. As little dietary methylmercury is

excreted in the urine, the contribution of ingested methylmercury to the measured levels is not expected

to be high. Good correlations have been observed between urinary mercury levels and air levels of

mercury vapor; however, these correlations were obtained after correcting urinary mercury content for

variations in the urinary excretion rate (using urinary creatinine content or specific gravity) and after

standardizing the amount of time elapsed after exposure (Roels et al. 1987). Such steps are necessary

because considerable intra- and interindividual variability has been observed in the urinary excretion rate

(Barber and Wallis 1986; Piotrowski et al. 1975). Even when such precautions are taken, intraindividual

variability remains at 18% (Barregard et al. 1992; Roels et al. 1987).

2.5.1.3 Exhaled Air

Exhaled air has been suggested as a possible biomarker of exposure to elemental mercury vapor

because a portion of absorbed mercury vapor is excreted via the lungs. Excretion by this route has a half-
d  CxbxV

Axf

2-11

life of approximately 18 hours (Hursh et al. 1976). At low levels of exposure, however, mercury vapor

released from dental amalgam may contribute substantially to the measured amount of mercury.

2.5.2 Inorganic Mercury

No information was identified in the literature that specifically assessed biological indicators for

inorganic mercury exposure. The information presented above for detection of mercury in blood and

urine after occupational exposure to elemental mercury vapor should also apply to inorganic mercury

exposures because elemental mercury vapor is rapidly converted to mercuric mercury after absorption.

2.5.3 Methylmercury

Blood and scalp hair are the primary indicators used to assess methylmercury exposure.

2.5.3.1 Blood

Because methylmercury freely distributes throughout the body, blood is a good indicator medium

for estimating methylmercury exposure. Because an individual's intake may fluctuate, blood levels may

not necessarily reflect mercury intake over time (Sherlock et al. 1982; Sherlock and Quinn, 1988). At

steady state, blood levels have been related to dose by the following equation (Kershaw et al. 1980):

Where:

C = concentration in blood (expressed in g/L)

V = volume of blood (expressed as L)

b = the kinetic rate constant (day )-1

A = absorption rate (unitless)

F = fraction of dose that is present in blood

d = intake (g/day)

It is useful to measure blood hematocrit and mercury concentrations in both whole blood and

plasma. From these data, the red blood cell to plasma mercury ratio may be determined, and interference

from exposure to high levels of elemental or inorganic mercury may be estimated (Clarkson et al. 1988).

2.5.3.2 Scalp Hair

Scalp hair can also be a good indicator medium for estimating methylmercury exposure (Phelps

et al. 1980). Methylmercury is incorporated into scalp hair at the hair follicle in proportion to its content

in blood. The hair-to-blood ratio in humans has been estimated as approximately 250:1 expressed as g

Hg/g hair to mg Hg/L blood, but some difficulties in measurements, interindividual variation in body

burden, differences in hair growth rates, and variations in fresh and saltwater fish intake have led to

varying estimates (Birke et al. 1972; Skerfving 1974). Once incorporated into the hair, the

methylmercury is stable, and, therefore, gives a longitudinal history of blood methylmercury levels
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(Phelps et al. 1980; WHO, 1990). Analysis of hair mercury levels may be confounded by adsorption of

mercury vapor onto the hair strands (Francis et al. 1982).

2.5.4 Methods of Analysis for Measuring Mercury in Biological Samples

The most common methods used to determine mercury levels in blood, urine and hair of humans

and animals include atomic absorption spectrometry (AAS), neutron activation analysis (NAA), X-ray

fluorescence (XRF) and gas chromatography (GC). Table 2-2 identifies the major characteristics of these

methods.

Table 2-2

Analytical Methods for the Detection of Mercury in Biological Samples

Method Methylmercury? Detection Limit (ppm) References

Able to Distinguish

NAA No 0.1 Byrne and Kosta (1974)

WHO (1976)

AAS No 2 Hatch and Ott (1968)

No PPB range Magos and Clarkson (1972) a

GC — Electron capture Yes 1.0 Von Burg et al. (1974)

Cappon and Smith (1978)

XRF No "low ppm" Marsh et al. (1987)

a

 The Magos and Clarkson method estimates methylmercury by subtracting the inorganic mercury content from

 the total mercury content.

2.6 Studies on Pharmacokinetic Models

2.6.1 Introduction

Pharmacokinetic modeling is a process by which administered dose, such as the amount of a

compound instilled into the body via inhalation, ingestion or parenteral route is used to estimate 

measures of tissue dose which may not always be accessible to measurement by direct experimentation. 

A pharmacokinetic model is employed to predict relevant measures of tissue dose under a wide range of

exposure conditions. In practice, the pharmacokinetic models used may incorporate features such as

compartmental analysis and physiologically-based models.

Reports available on the in vivo distribution of several types of mercury compounds provide

different physiokinetic relationships between the structure of mercury compounds and their behavior in

living organisms because the studies reported have been carried out under different experimental

conditions. Takeda et al. (1968) reported that in the rat, alkyl mercury compounds such as ethylmercuric

chloride and butylmercuric chloride were excreted more slowly and were retained in higher concentration

for a longer time in the body than mercuric chloride and phenylmercuric chloride. The distribution of

mercury in the brain was found to depend on the structure of the mercury compounds; relatively high

accumulation was observed for ethyl and n-butyl mercury compounds. Sebe and Itsuno (1962) reported

that after oral administration methyl-, ethyl-, and n-propylmercury compounds were neurotoxic to rats;
2-13

n-butylmercury was not neurotoxic and thus presumably did not cross the blood-brain barrier. By

contrast, Suzuki et al. (1963, 1964) reported that ethylmercuric acetate and n-butylmercuric acetate had

similar patterns of distribution when subcutaneously administered to mice.

2.6.2 Inorganic mercury

Few controlled laboratory studies of pharmacokinetics of mercury in humans have been

published (Hursh et al. 1976, Rahola et al. 1973). Rahola et al. (1973) examined mercury absorption and

elimination after oral administration of mercuric nitrate to five male and five female volunteers, and

reported very low and variable rate of gastrointestinal absorption (8 to 25% dose). They reported a halftime for inorganic mercury in human red blood cells of 16 days and whole body of 46 (32–60) days in

males and somewhat lower values in females. Hursh et al. (1976) found half-times for mercury clearance

from the body of 58 (35–90) days after exposure to mercury vapor. Whole body clearance from the

Rahola et al. (1973) study appeared biphasic with half-times of 2.3 days for the fast compartment and 42

(39–45) days for the slow compartment.

Low and variable rates of absorption of orally administered inorganic mercury in the Rahola et

al. (1973) study prompted Hall et al. (1994) to examine distribution of intravenously administered

inorganic mercury in human volunteers. In order to describe retention of mercury after transient

distributional effects, a one-compartment model was fit to the blood and body burden data after day 10,

assuming first order kinetics. The half-lives observed in the single compartment model for blood and

body burden were 30 (19.7–65.6 days) and 67 (48.6–95.5 days) days, respectively. The authors

attempted closer agreement between observed and predicted values by structuring a multicompartment

model. Measured mercury concentrations in blood, urine, feces, and whole body radioactive levels of

mercury were used in an a posteriori fashion to develop a model comprising six blood compartments,

one compartment each for feces and urine and a delayed compartment for feces. Inter-subject variability

(temporal pattern of blood mercury) and the existence of a kinetically distinct plasma pool (three distinct

compartments) for mercury resulted in equivocal predictions for blood, urine and feces; whether these

findings point to uncertainties of measurement of body burden or incomplete collection of excreta or

suggest other pathways of excretion, such as exhalation or sweating, is unknown. The authors concluded

that this type of complex pattern of blood kinetics, although unusual, is not without precedent. Four

kinetically distinct plasma pools of selenium has been reported after oral dosing with a stable isotopic

tracer (Patterson and Zech 1992). Hall et al. (1994) noted that the apparently linear kinetics observed for

the small tracer doses of i.v. inorganic mercury would likely change with toxicity associated with larger

or more frequent doses.

2.6.3 Methylmercury

Methylmercury is structurally the simplest of the organic mercurials; it bioaccumulates in certain

species of fish, some of which are important human and wildlife foods. In order to elucidate the

mechanisms that influence the pharmacokinetics of both methylmercury and mercuric mercury and to

extrapolate further both intra- and inter-species extrapolation of experimental data for these toxins, Farris

and associates (1993) developed a physiological pharmacokinetic model for methylmercury and its

metabolite, mercuric mercury. This was done in growing rats dosed orally with labeled methylmercury

over a period of 98 days. Mercuric mercury accounted for less than 0.5% of total activity. Extensive sets

of metabolism and distribution data were collected to understand the processes that influence the

pharmacokinetics of both methylmercury and mercuric mercury. The model consisted of nine lumped

compartments, each of which represented a major site of mercury accumulation, distribution or

elimination. The carcass served as a residual compartment, which included all tissues and organs not
2-14

separately incorporated into the model. Model simulations in this study were made with experimentally

determined concentrations of both inorganic and methylmercury in blood, brain, kidney and liver. The

data showed bidirectional and symmetric transport of both chemical species between blood and tissues

with relatively slow movement into and out of the brain. Some key parameters remained uncertain; for

example, the rate constant for demethylation is one of the most critical in adopting the model to other

species. This model, however, established a foundation for more complete understanding of

methylmercury pharmacokinetics. With further refinements, it could be applied to other species

including humans. To characterize health hazard from dietary methylmercury better, one needs to

understand the distribution of methylmercury in the body, the extent to which it accumulates and the rate

at which it is eliminated. Farris et al. (1993) noted that following methylmercury dosing there was a

buildup of inorganic mercury in tissues and that excreted mercury was predominantly mercuric;

methylmercury behaved as a single body pool, while mercuric mercury was handled differently in

different tissues.

Smith and associates (1994) made further refinements to the Farris et al. (1993) model. They

reported a multicompartment pharmacokinetic model for methylmercury and mercuric mercury in seven

human volunteers. This model simulated the long-term disposition of methylmercury and inorganic

mercury in humans following a single i.v. dose of radio-labeled methylmercury. This was a tracer

amount to avoid toxic or saturation effects. The behavior of both methylmercury and inorganic mercury

in the body was modeled with the simplest compartmental model which fitted the data; blood, urine and

feces data were used to fit the model. In this model the tracer dose was delivered to the first blood

compartment and subsequently distributed to two extra-vascular methylmercury compartments; two

distinct compartments (urine and feces) for inorganic mercury were added features. This fivecompartment model showed that inorganic mercury accumulated in the body and at longer times was the

predominant form of mercury present. The biological half-life of methylmercury in the body was

calculated to be 44 days, and 1.6% of the body burden was lost each day by both metabolism and

excretion.

To characterize neurological impairments of prenatal methylmercury exposure in children,

Gearhart and associates (1995) applied a more sophisticated multispecies pharmacokinetic model and

statistical dose-response analysis to an epidemiological study of a large population in New Zealand

(Kjellstrom et al. 1989) which featured relatively constant chronic exposure to methylmercury in fish. 

The model for methylmercury in this study consisted of an adult with 11 compartments representing both

organ-specific and lumped tissues; eight compartments represented transport of methylmercury as flowlimited, and three other compartments represented transport as diffusion-limited. The flow-limited

compartments were plasma, kidney, richly perfused, slowly perfused, brain-blood, placenta, liver and gut

compartments; RBC, brain and fetus were the diffusion-limited compartments. There were also four

other compartments in the model which were involved in methylmercury uptake and elimination: 

methylmercury in the urine; and methylmercury and inorganic mercury in the hair, feces and the

intestinal lumen. The fetal sub-model for methylmercury consisted of four compartments: fetal plasma,

RBCs, brain and the remaining fetal body. This modeling effort was designed to create a multispecies

model that would be amenable to simulation of the kinetics of methylmercury by simply changing the

species-specific parameters. Unlike Farris et al. (1993), separate red blood cell and plasma

compartments were used to predict changes in kinetics of methylmercury across species due to

differences in the red blood cell/plasma ratio. Different pharmacokinetic parameters, such as

tissue/blood partition coefficients and volume distributions for humans, rats and monkeys, were taken

from different studies published in the current literature. The authors provided a benchmark dose on

results of a battery of neurobehavioral tests in 6-year-old children prenatally exposed to methylmercury

in seafood. Their calculations suggested a NOAEL of 17 ppm Hg in maternal hair for the most sensitive
2-15

neurological event in children. The analysis of the pharmacokinetic model indicated that the fetal brain

concentrations of methylmercury at this NOAEL were on the order of 50 ppb and were associated with

maternal dietary intakes of methylmercury ranging from 0.8 to 2.5 g/kg-day. These analyses provided

support to the Iraqi data used in the development of the RfD for methylmercury, presented in the risk

assessment chapter (Chapter 6) of this volume.

2.6.4 Discussion

Both simple and complex multi-compartment models have been reported by Hall et al. (1994),

Farris et al. (1993), Smith et al. (1994) and Gearhart et al. (1995). The Hall et al. (1994) paper discussed

a model which employed inorganic mercury data obtained from human studies; however, temporal

patterns of blood mercury and the existence of kinetically distinct plasma pools for mercury present

uncertainties which limit the use of this model in risk assessment. Farris et al. (1993) reported a

multicompartment model using data obtained from rats exposed to methylmercury in diets over a period

of 98 days. They observed a buildup of inorganic mercury in tissues and the conversion of

methylmercury to inorganic mercury could not be accurately predicted by whole-body counting, which

was also subjected to errors from low sensitivity and the inability to compensate for geometric changes

due to redistribution of methylmercury or translocation of inorganic mercury to its target tissues. Smith

and associates (1994) refined this model and presented a multicompartment model using data obtained

from humans given a single i.v. dose of methylmercury. Uncertainties, however, persist in prediction of

methylmercury exposures in food. Since methylmercury causes subtle neurotoxicity in children, this

model may not be predictive of exposure in children. This potential neurotoxicity observed in prenatally

exposed children prompted Gearhart et al. (1995) to develop multicompartment adult and fetal model

using data from rat, monkey and humans. This model was applied to an epidemiology study on which

benchmark dose analysis was used to better characterize the dose-response information rather than the

traditional NOAEL approach. In the risk assessment chapter of this volume, U.S. EPA utilizes a

benchmark dose approach for setting the RfD for methylmercury. A multispecies compartment model

discussed in the Gearhart et al. (1995) report may provide a viable approach because it can use data from

both adults and neonates. This approach can use adult and neonatal effects data from several animal and

human studies to account for evidences of non-linearities in dose-responses. Research is needed to

reduce uncertainties in racial, ethnic, and cultural differences which exist in epidemiological studies.
3-1

3. BIOLOGICAL EFFECTS

This chapter summarizes the available toxicity data on mercury compounds; the information is

tabulated specifically for each form of mercury (i.e., elemental, inorganic and methylmercury) and each

toxicity endpoint. Case studies in humans are distinguished from epidemiological studies and animal

studies. In addition, critical studies for a given endpoint are briefly summarized in the narrative

preceding the corresponding table.

The tables provide information on study design, observed effects, study limitations and any

reported biological monitoring levels (BMLs) of mercury. To the extent possible, BML values have been

reported in consistent units throughout this chapter (μg/L in body fluids, μg/g in tissue, μg/g creatinine in

urine). It was not possible, however, to use completely consistent units because investigators measured

mercury in different media (e.g., blood, urine, or tissue) or used different time frames (e.g., μg/L urine,

μg/24 hour urine). In addition, some investigators normalized urine concentrations to the amount of

creatinine, while most did not. An explanation is provided in Appendix A for any dose conversions

required during review and evaluation of the toxicity and carcinogenicity studies reported in the

discussions presented below.

3.1 Elemental Mercury

3.1.1 Critical Noncancer Data

This section describes studies evaluated by U.S. EPA for use in assessing general systemic health

risks, primarily toxicity in exposed workers. Chapter 6 describes the derivation of an inhalation

Reference Concentration (RfC) for elemental mercury based on neurotoxicity observed in several human

occupational studies. For completeness, some of these studies are also presented in tabular form in

succeeding sections.

Fawer et al. (1983) used a sensitive measure of intention tremor (tremors that occur at the

initiation of voluntary movements) in workers occupationally exposed for an average of 26 years to

metallic mercury vapor. A statistically significant difference was seen in the frequency of these tremors

in mercury-exposed workers compared with unexposed workers. The concentration of metallic mercury

in the air was measured, and a time-weighted-average (TWA) of 0.026 mg/m over an average of 15.3 3

years was derived. This was based on the assumption that the workers were exposed to the same

concentration of mercury for the duration of their employment. It should be noted that very little detail

was presented with regard to the measurement of the exposure levels, and that it is likely that there were

variations in the mercury air levels during the period of exposure. Furthermore, the tremors may have

resulted from intermittent exposure to concentrations higher than the TWA.

Piikivi and Tolonen (1989) studied the effects of long-term exposure to mercury vapor on the

electroencephalograms (EEG) of 41 chloralkali workers exposed for a mean of 15.6 ± 8.9 years by

comparison to matched referent controls. They found that the exposed workers, who had mean blood

mercury levels of 12 g/L and mean urine mercury levels of 20 g/L, tended to have an increased

number of EEG abnormalities when analyzed by visual inspection only. When the EEGs were analyzed

by computer, the exposed workers had significantly slower and attenuated EEGs as compared to the

referents. These changes were observed in 15% of the exposed workers. The frequency of these changes

correlated with cortical mercury content (measured in other studies); the changes were most prominent in

the occipital cortex, less prominent in the parietal cortex and almost absent in the frontal cortex. The
3-2

authors extrapolated an exposure level associated with these EEG changes of 0.025 mg/m from blood 3

levels based on a conversion factor calculated by Roels et al. (1987).

Piikivi and Hanninen (1989) studied the subjective symptoms and psychological performances

on a computer-administered test battery in 60 chloralkali workers exposed to mercury vapor for a mean

of 13.7 ± 5.5 years as compared to matched referent controls. The exposed workers had mean blood

mercury levels of 10 g/L and mean urine mercury levels of 17 g/L. A statistically significant increase

in subjective measures of memory disturbance and sleep disorders was found in the exposed workers. 

The exposed workers also reported more anger, fatigue and confusion. No objective disturbances in

perceptual motor, memory or learning abilities were found in the exposed workers. The authors

extrapolated an exposure level associated with these subjective measures of memory disturbance of 0.025

mg/m from blood levels based on a conversion factor calculated by Roels et al. (1987). 3

Both subjective and objective symptoms of autonomic dysfunction were investigated in

41 chloralkali workers exposed to mercury vapor for a mean of 15.6 ± 8.9 years as compared to matched

referent controls (Piikivi 1989). The quantitative non-invasive test battery consisted of measurements of

pulse rate variation in normal and deep breathing in the Valsalva maneuver and in vertical tilt, as well as

blood pressure responses during standing and isometric work. The exposed workers had mean blood

levels of 11.6 g/L and mean urinary levels of 19.3 g/L. The exposed workers complained of more

subjective symptoms of autonomic dysfunction than the controls, but the only statistically significant

difference was an increased reporting of palpitations in the exposed workers. The quantitative tests

revealed a slight decrease in pulse rate variations, indicative of autonomic reflex dysfunction, in the

exposed workers. The authors extrapolated an exposure level associated with these subjective and

objective measures of autonomic dysfunction of 0.03 mg/m from blood levels based on the conversion 3

factor calculated by Roels et al. (1987).

Sensory and motor nerve conduction velocities were studied in 18 workers from a mercury cell

chlorine plant (Levine 1982). Time-integrated urine mercury levels were used as an indicator of mercury

exposure. Using linearized regression analysis, the authors found that motor and sensory nerve

conduction velocity changes, (i.e., prolonged distal latencies) were correlated with the time-integrated

urinary mercury levels in asymptomatic exposed workers and occurred when urinary mercury levels

exceeded 25 g/L. This study demonstrated that elemental mercury exposure can be associated with

preclinical evidence of peripheral neurotoxicity.

Miller et al. (1975) investigated several subclinical parameters of neurological dysfunction in

142 workers exposed to inorganic mercury in either the chloralkali industry or a factory for the

manufacture of magnetic materials. They found that there was a significant increase in average forearm

tremor frequency in workers whose urinary mercury concentration exceeded 50 g/L as compared to

unexposed controls. Also observed were eyelid fasciculation, hyperactive deep tendon reflexes and

dermatographia, but there was no correlation between the incidence of these findings and urinary

mercury levels.

Roels et al. (1985) examined 131 male and 54 female workers occupationally exposed to

mercury vapor for an average duration of 4.8 years. Urinary mercury (52 and 37 g/g creatinine for

males and females, respectively) and blood mercury levels (14 and 9 g/L for males and females,

respectively) were recorded, but atmospheric mercury concentration was not provided. Symptoms

indicative of central nervous system (CNS) disorders were reported but were not related to mercury

exposure. Minor renal tubular effects were detected in mercury-exposed males and females and were

attributed to current exposure intensity rather than exposure duration. Male subjects with urinary
3-3

mercury levels of >50 g/g creatinine exhibited preclinical signs of hand tremor. It was noted that

females did not exhibit this effect, and that their urinary mercury never reached the level of 50 g/g

creatinine. A companion study (Roels et al. 1987) related air mercury levels to blood mercury

(Hgblood) and urinary mercury (HgU) values in 10 workers in a chloralkali battery plant. Duration of

exposure was not specified. A high correlation was reported for Hgair and HgU for pre-shift exposure

(r=0.70, p<0.001) and post-shift (r=0.81, p<0.001) measurements. Based on these data and the results of

their earlier (1985) study, the investigators suggested that some mercury-induced effects may occur when

HgU levels exceed 50 g/g creatinine and that this value corresponds to a mercury TWA of 0.04 mg

Hg/m .3

A survey of 567 workers at 21 chloralkali plants was conducted to ascertain the effects of

mercury vapor inhalation (Smith et al. 1970). Mercury levels ranged from <0.01 to 0.27 mg/m , and 3

chlorine concentrations ranged from 0.1 to 0.3 ppm in most of the working stations of these plants. 

Worker exposure to mercury levels (TWA) varied with 10.2% of the workers exposed to <0.01 mg

Hg/m , 48.7% exposed to 0.01–0.05 mg Hg/m , 25.6% exposed to 0.06–0.10 mg Hg/m , and 4.8% 333

exposed to 0.24–0.27 mg Hg/m (approximately 85% were exposed to mercury levels 0.1 mg/m ). The 3 3

duration of employment for the examined workers ranged from one year (13.3%) to >10 years (31%)

with 55.7% of the workers employed for 2 or 9 years. A group of 600 workers not exposed to chlorine

served as a control group. A strong positive correlation (p<0.001) was found between the mercury

TWAs and the reporting of subjective neuropsychiatric symptoms (nervousness, insomnia), occurrence

of objective tremors and weight and appetite loss. A positive correlation (p<0.001) was also found

between mercury exposure levels and urinary and blood mercury levels of test subjects. No adverse

alterations in cardiorespiratory, gastrointestinal, renal, or hepatic functions were attributed to the mercury

vapor exposure. Additionally, biochemical (hematologic data, enzyme activities) and clinical

measurements (electrocardiogram, chest X-rays) were not different between the mercury-exposed and

non-exposed workers. No significant signs or symptoms were noted for individuals exposed to mercury

vapor concentrations 0.1 mg Hg/m . This study provides data indicative of a no-observed-adverse- 3

effect level (NOAEL) of 0.1 mg Hg/m and a lowest-observed-adverse-effect level (LOAEL) of 0.18 mg 3

Hg/m . In a follow-up study conducted by Bunn et al. (1986), however, no significant differences in the 3

frequency of objective or subjective findings such as weight loss and appetite loss were seen in workers

exposed to mercury at levels that ranged between 50 and 100 mg/m . The study by Bunn et al. (1986) 3

was limited, in that little information was provided regarding several methodological questions such as

quality assurance measures and control of possible confounding variables.

Neurological signs and symptoms (i.e., tremors) were observed in 79 workers exposed to metallic

mercury vapor; urinary mercury levels in affected subjects exceeded 500 g/L. Short-term memory

deficits were seen in workers whose urine levels were less than 500 g/L (Langolf et al. 1978). Impaired

performance in mechanical and visual memory tasks and psychomotor ability tests was reported by Forzi

et al. (1978) in exposed workers whose urinary mercury levels exceeded 100 g/L.

Decreased strength, decreased coordination, increased tremor, decreased sensation and increased

prevalence of Babinski and snout reflexes were exhibited by 247 exposed workers whose urinary

mercury levels exceeded 600 g/L. Evidence of clinical neuropathy was observed at urinary mercury

levels that exceeded 850 g/L (Albers et al. 1988). Preclinical psychomotor dysfunction was reported to

occur at a higher incidence in 43 exposed workers (mean exposure duration of 5 years) whose mean

urinary excretion of mercury was 50 g/L. In the same study, workers whose mean urinary mercury

excretion was 71 g/L had a higher incidence of total proteinuria and albuminuria (Roels et al. 1982). 

Postural and intention tremor was observed in 54 exposed workers (mean exposure duration of 7.7 years)

whose mean urinary excretion of mercury was 63 g/L (Roels et al. 1989). Verbeck et al. (1986)
3-4

observed an increase in tremor parameters with increasing urinary excretion of mercury in 21 workers

exposed to mercury vapor for 0.5–19 years. The LOAEL for this effect was a mean urinary excretion of

35 g/g creatinine.

The elemental mercury levels reported to be associated with preclinical and symptomatic

neurological dysfunction are generally lower than those found to affect kidney function, as discussed

below.

Piikivi and Ruokonen (1989) found no evidence of glomerular or tubular damage in

60 chloralkali workers exposed to mercury vapor for an average of 13.7 ± 5.5 years as compared to their

matched referent controls. Renal function was assessed by measuring urinary albumin and N-acetyl--

glucosaminidase (NAG) activity. The mean blood mercury level in the exposed workers was 14 g/L,

and the mean urinary mercury level was 17 g/L. The authors extrapolated a NOAEL for kidney effects

based on these results of 0.025 mg/m from blood levels based on the conversion factor calculated by 3

Roels et al. (1987).

Stewart et al. (1977) studied urinary protein excretion in 21 laboratory workers exposed to

0.01–0.05 mg/m of mercury. Their urinary level of mercury was ~35 g/L. Increased proteinuria was 3

found in the exposed workers by comparison to unexposed controls. When preventive measures were

instituted to limit exposure to mercury, proteinuria was no longer observed in the exposed technicians.

Lauwerys et al. (1983) found no change in several indices of renal function (e.g., proteinuria,

albuminuria, urinary excretion of retinol-binding protein, aminoaciduria, creatinine in serum,

2-microglobulin in serum) in 62 workers exposed to mercury vapor for an average of 5.5 years. The

mean urinary mercury excretion in the exposed workers was 56 g/g creatinine, which corresponds to an

exposure level of 0.046 mg/m according to a conversion factor of 1:1.22 (air:urine [g/g creatinine]) 3

(Roels et al. 1987). Despite the lack of renal effects observed, 8 workers were found to have an increase

in serum anti-laminin antibodies, which can be indicative of immunological effects. In a follow-up study

conducted by Bernard et al. (1987), however, there was no evidence of increased serum anti-laminin

antibodies in 58 workers exposed to mercury vapor for an average of 7.9 years. These workers had a

mean urinary mercury excretion of 72 g/g creatinine, which corresponds to an exposure level of 0.059

mg/m .3

Renal function in 100 chloralkali workers exposed to inorganic mercury vapor for an average of

8 years was studied (Stonard et al. 1983). No changes in the following urinary parameters of renal

function were observed at mean urinary mercury excretion rates of 67 g/g creatinine: total protein,

albumin, 1 2 -acid glycoprotein,  -microglobulin, NAG and -glutamyl transferase. When urinary

mercury excretion exceeded 100 g/g creatinine, a small increase in the prevalence of higher activities of

NAG and -glutamyl transferase were observed.

Rosenman et al. (1986) evaluated routine clinical parameters (physical exams, blood chemistry,

urinalysis), neuropsychological disorders, urinary NAG, motor nerve conduction velocities and

occurrence of lenticular opacities in 42 workers of a chemical plant producing mercury compounds. A

positive correlation (p<0.05 to p<0.001) was noted between urinary mercury (levels ranged from

100–250 g/L) and the number of neuropsychological symptoms, NAG excretions and decrease in motor

nerve conduction velocities. Evidence of renal dysfunction was seen in 63 chloralkali workers. This

included increased plasma and urinary concentrations of -galactosidase, increased urinary excretion of

high-molecular weight proteins and a slightly increased plasma 2-microglobulin concentration. The
3-5

incidence of these effects increased in workers whose urinary mercury excretion exceeded 50 g/g

creatinine (Buchet et al. 1980).

Increased urinary NAG levels were found in workers whose urinary mercury levels exceeded

50 g/L (Langworth et al. 1987). An increase in the concentration of urinary brush border proteins (BB50) was observed in 20 workers whose mean urinary mercury excretion exceeded 50 g/g creatinine

(Mutti et al. 1985). Foa et al. (1976) found that 15 out of 81 chloralkali workers exposed to 0.06–0.30

mg/m mercury exhibited proteinuria. An increased excretion of -glutamyl transpeptidase, indicative of 3

renal dysfunction, was found in 509 infants dermally exposed to phenylmercury via contaminated diapers

(Gotelli et al. 1985).

The elemental mercury levels reported to be associated with preclinical and symptomatic

neurological dysfunction and kidney effects are lower than those found to affect pulmonary function, as

discussed below.

McFarland and Reigel (1978) described the cases of six workers who were acutely exposed (4–8

hours) to calculated metallic mercury vapor levels of 1.1–44 mg/m . These men exhibited a combination 3

of chest pains, dyspnea, cough, hemoptysis, impairment of pulmonary function (reduced vital capacity),

diffuse pulmonary infiltrates and evidence of interstitial pneumonitis. Although the respiratory

symptoms resolved, all six cases exhibited chronic neurological dysfunction, presumably as a result of

the acute, high-level exposure to mercury vapor.

Lilis et al. (1985) described the case of a 31-year-old male who was acutely exposed to high

levels of mercury vapor in a gold extracting facility. Upon admission to the hospital, the patient

exhibited dyspnea, chest pain with deep inspiration, irregular infiltrates in the lungs and reduced

pulmonary function (forced vital capacity). The level of mercury to which he was exposed is not known,

but a 24-hour urine collection contained 1,900 g Hg/L. Although the patient improved gradually over

the next several days, he still showed signs of pulmonary function abnormalities (e.g., restriction and

diffusion impairment) 11 months after exposure.

Levin et al. (1988) described four cases of acute high-level mercury exposure during gold ore

purification. The respiratory symptoms observed in these four cases ranged from minimal shortness of

breath and cough to severe hypoxemia. The most severely affected patient exhibited mild interstitial lung

disease both radiographically and on pulmonary function testing. One patient had a urinary mercury

level of 245 g/L upon hospital admission. The occurrence of long-term respiratory effects in these

patients could not be evaluated since all but one refused follow-up treatment.

Schweinsberg (1994) evaluated mercury in hair, blood and urine of subjects who had amalgam

fillings, who consumed fish or who had occupational exposure. The first group consisted of 67 males

aged 16-72 yrs, mean age 36. The fish-eating population was 149 males and females (age 22-80, mean =

47) who either did or did not eat fish from the Rhine River. The workers were 105 male and female

employees of a Thuringian thermometer plant (ages 19- 65, mean = 42). Both fish consumption and the

presence of mercury amalgam fillings resulted in measurable mercury in blood or urine. The range of

mercury in the workers was higher by about 100 fold. The authors present most of the data graphically;

they report that persons without amalgam fillings who did not consume Rhine fish or work in the

thermometer factory had blood mercury in the range of 0.2 μg/l (detection limit) and 0.4 μg/l (n not

given). It appears that from Figure 5 of the paper that blood mercury levels ranged between 100-120 

μg/l in 5 workers. The authors state that no changes in nerve conduction velocities or N-acetyl--Dglucosaminidase activities were observed (data not shown in paper). 
3-6

3.1.2 Cancer Data

3.1.2.1 Human Data

A number of epidemiological studies were conducted that examined mortality among elemental

mercury vapor-exposed workers. Conflicting data regarding a correlation between mercury exposure and

an increased incidence of cancer mortalities have been obtained. All of the studies have limitations

which compromise interpretation of their results. These studies are summarized in Table 3-1.

A retrospective cohort study examined mortality among 5,663 white males who worked between

1953 and 1958 at a plant in Oak Ridge, Tennessee, where elemental mercury was used for lithium isotope

separation (Cragle et al. 1984). The workers were divided into three cohorts: exposed workers who had

been monitored on a quarterly basis for mercury levels in urine (n=2,133); workers exposed in the

mercury process section for whom urinalysis monitoring data were not collected (n=270); and unexposed

workers from other sections of the nuclear weapons production facility (n=3,260). The study subjects

worked at least 4 months during 1953–1958 (a period when mercury exposures were likely to be high);

mortality data from death certificates were followed through the end of 1978. The mean age of the men

at first employment at the facility was 33 years, and average length of their employment was greater than

16 years with a mean 3.73 years of estimated mercury exposure. Air mercury levels were monitored

beginning in 1955, and during 1955 through the third quarter of 1956, air mercury levels were above 100

μg/m in 30–80% of the samples. Thereafter, air mercury levels decreased to concentrations below 100 3

μg/m . The mortality experience [standard mortality ratio (SMR)] of each group was compared with the 3

age-adjusted mortality experience of the U.S. white male population. Among exposed and monitored

workers, there were no significant increases in mortality from cancer at any site, even after the level or

length of exposure was considered. A significantly lower mortality from all causes was observed. There

was an excess of deaths due to lung cancer in the exposed, monitored workers (42 observed, 31.36

expected) but also in the unexposed workers (71 observed and 52.93 expected). The SMR for each group

was 1.34; the elevated incidence of lung cancer deaths was, therefore, attributed to some other factor at

the plant and/or to lifestyle factors (e.g., smoking) common to both the exposed and unexposed groups. 

Study limitations include small cell sizes for cancer mortality, which limited the statistical stability of

many comparisons.

Barregard et al. (1990) studied mortality and cancer morbidity between 1958 and 1984 in 1,190

workers from eight Swedish chloralkali plants that used the mercury cell process in the production of

chlorine. The men included in the study had been monitored for urinary or blood mercury for more than

one year between 1946 and 1984. Vital status and cause of death were ascertained from the National

Population Register and the National Bureau of Statistics. The cancer incidence of the cohort was

obtained from the Swedish Cancer Register. The observed total mortality and cancer incidences were

compared with those of the general Swedish male population. Comparisons were not made between

exposed and unexposed workers. Mean urinary mercury levels indicated a decrease in exposure between

the 1950s and 1970s; the mean urinary mercury level was 200 μg/L during the 1950s, 150 μg/L during

the 1960s and 50 μg/L in the 1970s. Mortality from all causes was not significantly increased in exposed

workers. A significant increase in deaths from lung tumors with greater than 10 years of latency was

seen in exposed workers (rate ratio, 2.0; 95% C.I. 1.0–3.8), but 9 of the 10 observed cases of lung cancer

occurred among workers (457 of the 1,190) possibly exposed to asbestos as well as to mercury. No dose

response was observed with respect to mercury exposure and lung tumors. This study is limited because

no quantitation was provided on smoking status, and results were confounded by exposure to asbestos.
3-7

Table 3-1

Carcinogenic Effects of Elemental Mercury in Humans: Epidemiological Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/ 4 months NS, but up No biologically significant increase in cancer mortality in workers Cragle et al. 1984

2,133 M (occup) to 80% of at an isotope enrichment plant, compared with unexposed workers

air samples at the same plant, or with age-adjusted mortality of U.S. males. 

in early Lung cancer mortality was increased in exposed workers, but the

years were increase was not statistically significant and a significant increase

>0.10; this in lung cancer mortality was observed in unexposed workers.

declined to Limitations: small cell size for cancer mortality, limiting

1-10% in statistical power of comparisons

later years BML not reported

Human/376 NS NS Increased lung cancer incidence among female hat makers Buiatti et al. 1985

cases of lung (occup) (p = 0.01). Controls were matched by age, sex and smoking

cancer (6 were history.

hatmakers), 892 Limitations: Hat makers were also exposed to arsenic

controls BML not reported

Human/ NS NS Increased incidence of glioblastomas among dental professionals. Ahlbom et al. 1986

cohort/3454 M, (occup) 95% C.I. = 1.3-3.4. Expected incidence was based on all

5787 F employed people (Sweden), stratified by age, sex and county. 

Limitations: No information was provided on the duration or level

of exposure; subjects were also exposed to chloroform and X-rays.

BML not reported

Human/9,912 M NS NS Increased lung cancer mortality among metal miners (95% Amandus and

(369 silicotics, (occup) C.I. = 0.94-2.90 for silicotics, 0.98-1.42 for nonsilicotics). Costello 1991

9,543 Limitations: Miners were exposed to a variety of metals. Only

nonsilicotics) 274 worked in mercury mines, and data were not reported

separately for this group. Workers were also exposed to radon,

increase may have been related to silicosis 

BML not reported

Human/ 1 year NS Increased incidence of lung cancer among chloralkali workers, but Ellingsen et al. 1992

cohort/799 M (occup) there was no association with cumulative mercury dose, years of

employment or latency (95% C.I. = 1.0-2.6). The increase could

be partly explained by an assumed higher smoking incidence and

asbestos exposure. Cancer mortality and incidence compared with

age-adjusted Norwegian population.

Limitations: Subjects were also exposed to chlorine and low levels

of asbestos dust. Limited data available, since reported as an

abstract.

BML not reported

Human/1190 M at least one NS Cancer and mortality rates compared with general population; no Barregard et al. 1990

year grouped by increase in mortality; excess of lung cancers (rate ratio = 2.0; 95%

years x U- CI 1.0-3.8)

Hg; Limitations: co-exposure to asbestos

 >1000

ug/L,

1000-2000

ug/L,

>2000 ug/L
3-8

Ahlbom et al. (1986) examined the cancer mortality during 1961 to 1979 of cohorts of Swedish

dentists and dental nurses aged 20–64 and employed in 1960 (3,454 male dentists, 1,125 female dentists,

4,662 female dental nurses). Observed incidences were compared with those expected based on cancer

incidence during 1961–1979 among all Swedes employed during 1960 and the proportion of all Swedes

employed as dentists and dental nurses. Data were stratified by sex, age (5-year age groups), and county. 

The incidence of glioblastomas among the dentists and dental nurses combined was significantly

increased (SMR, 2.1; 95% C.I. 1.3–3.4); the individual groups had elevated SMRs (2.0–2.5), but the 95%

confidence intervals of these groups included unity. By contrast, physicians and nurses had SMRs of

only 1.3 and 1.2, respectively. Exposure to mercury could not be established as the causative factor

because exposure to other chemicals and X-rays was not ruled out.

Amandus and Costello (1991) examined the association between silicosis and lung cancer

mortality between 1959 and 1975 in white male metal miners (n=9,912) employed in the United States

between 1959 and 1961. Mercury exposures were not monitored. Exposures to specific metals among

the silicotic and nonsilicotic groups were analyzed separately. Lung cancer mortality in both silicotic

and nonsilicotic groups was compared with rates in white males in the U.S. population. Both silicotic

(n=11) and nonsilicotic mercury miners (n=263) had significantly increased lung cancer mortality (SMR,

14.03, 95% C.I., 2.89–40.99 for silicotics; SMR, 2.66, 95% C.I. 1.15–5.24 for nonsilicotics). The

analysis did not focus on mercury miners, and confounders such as smoking and radon exposure were not

analyzed with respect to mercury exposure. This study is also limited by the small sample size for

mercury miners.

A case-control study of persons admitted to a hospital in Florence, Italy with lung cancer

between 1981-1983 was performed to evaluate occupational risk factors (Buiatti et al. 1985). Cases were

matched with one or two controls (persons admitted to the hospital with diagnoses other than lung cancer

or suicide) with respect to sex, age, date of admission, and smoking status. Women who had "ever

worked" as hat makers had a significantly increased risk of lung cancer (p=0.01; determined using the

Mantel-Haenszel Chi-square test). The duration of employment as a hat maker averaged 22.2 years, and

latency averaged 47.8 years. Workers in the Italian hat industry were known to be occupationally

exposed to mercury; however, the design of this study did not allow evaluation of the relationship

between cumulative exposure and cancer incidence. In addition, interpretation of the results of this study

is limited by the small sample size (only 6/376 cases reported this occupation) and by exposure of hat

makers to other pollutants including arsenic, a known lung carcinogen.

Ellingsen et al. (1992) examined the total mortality and cancer incidence among 799 workers

employed for more than 1 year in two Norwegian chloralkali plants. Mortality incidence between 1953

and 1988 and cancer incidence between 1953 and 1989 were examined. Mortality and cancer incidence

were compared with that of the age-adjusted general male Norwegian population. No increase in total

cancer incidence was reported, but lung cancer was significantly elevated in the workers (ratio, 1.66;

95% C.I. 1.0–2.6). No causal relationship can be drawn between mercury exposure and lung cancer

because no correlation existed between cumulative mercury dose, years of employment, or latency time. 

Also, the prevalence of smoking was 10–20% higher in the exposed workers and many workers were also

exposed to asbestos.

3.1.2.2 Animal Data

Druckrey et al. (1957) injected 0.1 mL of metallic mercury intraperitoneally into 39 rats (males

and females; numbers of each not specified) of the BD III and BD IV strains. Among the rats surviving

longer than 22 months, 5 out of 12 developed peritoneal sarcomas (three females and two males). All
3-9

sarcomas were observed to have droplets of mercury present. Although severe kidney damage was

reported in all treated animals, there were no renal tumors or tumors at any site other than the peritoneal

cavity.

3.1.3 Other Data

3.1.3.1 Death

Accidental exposure to high concentrations of elemental mercury vapor for short amounts of time

has led to deaths in humans (Table 3-2). The cause of death in all available reports was respiratory

failure. The onset of death occurred six hours to 23 days after exposure to mercury vapors (Campbell

1948; Kanluen and Gottlieb 1991; Rowens et al. 1991; Soni et al. 1992; Taueg et al. 1992). Urinary

mercury concentrations indicated that body levels were up to 10 times higher than controls. Only acuteduration studies were found that directly linked elemental mercury vapor exposure to death.

Table 3-2

Lethality of Elemental Mercury in Humans: Case Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/1 F 5 hr NS Increased creatinine excretion; necrotic stomach mucosa; Campbell 1948

(4-month old) degeneration of convoluted tubules; death due to pulmonary

edema

Limitation: Limited exposure data

BML not reported

Human/2 M, 24 hr NS Respiratory distress; CNS alterations; nausea; tubular Kanluen and

2 F (adult, necrosis of proximal tubules in kidneys Gottlieb 1991;

2 elderly) Limitation: Limited exposure data Rowens et al. 1991

BML Range: 4.6-219 μg/L in urine

Human/1 F <6 hr NS Breathing difficulty; distended abdomen Soni et al. 1992

(1-yr old) Limitation: Limited exposure data

BML not reported

Human/2 M, NS 0.91 at 11- Dyspnea; respiratory failure; death at 11-24 days Taueg et al. 1992

2 F (adults) (Acute) 18 days post- postexposure

exposure BML Range: 94-423 μg/L in urine

Human/2 Several 0.01-0.04 Numbness in fingers and toes; absence of deep tendon Taueg et al. 1992

F (children) months several reflexes; visual field defects; weakness

months after BML not reported

initial spill

Animal studies reveal that pulmonary edema and asphyxiation result from acute highconcentration exposure to elemental mercury vapors (Table 3-3). Exposure to elemental mercury vapors

for two hours at a concentration of 27 mg Hg/m resulted in death of 20 of 32 rats (Livardjani et al. 3

1991). Rabbits exposed for 1 to 30 hours to 28.8 mg Hg/m of elemental mercury vapors appeared to be 3
3-10

less affected. Death occurred in only one of two rabbits exposed for 30 hours (Ashe et al. 1953). 

Exposure to the same concentration for a shorter duration resulted in no deaths.

Table 3-3

Lethality of Elemental Mercury in Animals: Inhalation Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Rabbit/strain 1-30 hr 28.8 LD for 30 hours; all rabbits exposed for shorter periods Ashe et al. 1953

NS/14 (sex survived.

NS) Limitations: There was no control group, and details on effects

50

were lacking.

BML: 5,320 μg/L in blood

Rat/Wistar/64 1 or 2 hr 0, 30 Death was due to asphyxiation; pulmonary edema and fibrosis Livardjani et al.

M/duration were observed. No animals exposed for 1 hour died by 15 days, 1991

and all animals exposed for 2 hours died within 5 days.

Limitations: No control group; limited reporting of histology

BML Range: 391-4,558 μg/L in blood at 1-15 days

postexposure

3.1.3.2 Neurological

Case reports from accidental exposures to high concentrations of mercury vapors (Adams et al.

1983; Aronow et al. 1990; Barber 1978; Bluhm et al. 1992a; Fagala and Wigg 1992; Foulds et al. 1987;

Friberg et al. 1953; Hallee 1969; Jaffe et al. 1983; Karpathios et al. 1991; Lilis et al. 1985; McFarland

and Reigel 1978; Sexton et al. 1976; Snodgrass et al. 1981; Taueg et al. 1992) as well as studies of

populations chronically exposed to potentially high concentrations (Ehrenberg et al. 1991; Friberg et al.

1953; Roels et al. 1982; Sexton et al. 1978) have provided considerable information about the

neurotoxicity of elemental mercury vapor. These studies have shown effects on a wide variety of

cognitive, sensory, personality and motor functions.

Occasionally, hearing loss, visual disturbances (visual field constriction), and/or hallucinations

have also occurred. In general, symptoms have been observed to subside after removal from exposure. 

However, persistent effects (tremor, cognitive deficits) have been observed in occupationally exposed

subjects 10 to 20 years after cessation of exposure (Albers et al. 1988; Ellingsen et al. 1993; Kishi et al.

1993).
3-11

Symptoms of Mercury Vapor-induced Neurotoxicity

The most prominent symptoms associated with mercury vapor-induced neurotoxicity include the following:

 tremors -- initially affecting the hands and sometimes spreading to other parts of the body

 emotional lability -- often referred to as "erethism" and characterized by irritability, excessive 

shyness, confidence loss and nervousness

 insomnia

 neuromuscular changes -- weakness, muscle atrophy, muscle twitching

 headaches

 polyneuropathy -- paresthesias, stocking-glove sensory loss, hyperactive tendon reflexes, slowed

sensory and motor nerve conduction velocities

 memory loss and performance deficits in tests of cognitive function

Studies of workers exposed to elemental mercury vapor have reported frank neurotoxicity at

exposure levels greater than 0.1 mg/m (Smith et al. 1970) or at levels resulting in urinary mercury of 3

greater than 300 μg in a 24-hour urine sample (Bidstrup et al. 1951). Several studies, however, have

shown evidence of neurotoxicity at approximately 2- to 4-fold lower concentrations. Self-reported

memory disturbances, sleep disorders, anger, fatigue, confusion and/or hand tremors were increased in

workers chronically exposed to an estimated 0.025 mg/m (blood levels of approximately 10 μg/L) 3

(Langworth et al. 1992a; Piikivi and Hanninen 1989). Also, objective measures of cognitive and/or

motor function in exposed populations have shown significant differences from unexposed controls

(Ehrenberg et al. 1991; Fawer et al. 1983; Liang et al. 1993; Ngim et al. 1992; Piikivi and Tolonen 1989;

Piikivi et al. 1984; Roels et al. 1982, 1989).

Table 3-4

Neurotoxicity of Elemental Mercury in Humans: Case Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/1 M 8-9 mo 0.02-0.45 Fatigue, irritability in an electrochemical industry worker Friberg et al. 1953

(adult) (occup) Limitations: small sample size; concomitant exposure to

chlorine; limited data reporting

BML: 680-900 μg/L in urine

Human/6 M <8 hr 44.3 Tremor; irritability; visual and hearing abnormalities McFarland and

(est.) Limitations: small sample size; limited data reporting Reigel 1978

BML Range: 1,060-3,280 μg/24 hr urine

Human/5 M, 6 51-176 d 0.1-1.0 Nervousness, insomnia and inattentiveness were more Sexton et al. 1978

F (adults and common than in controls; altered EEGs and personality

children)/ changes also noted

12 controls Limitations: small sample size

(sex NS) BML: 183-620 μg/L in blood at first measure

Human/2 M, 2 3 d NS Headache, slowed speech Snodgrass et al. 1981

F (adults) Limitation: Small sample size; limited exposure data

BML: 82-5700 μg/24 hr urine
Table 3-4 (continued)

Neurotoxicity of Elemental Mercury in Humans: Case Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

3-12

Human/1 M 2 d NS Delayed neurotoxicity: paresthesias; muscle Adams et al. 1983

(adult) fasciculations; hyperactive deep tendon reflexes

Limitation: small sample size; exposure data limited

BML: 98.75 μg/L in urine 3.5 months after exposure

Human/1 F (8- 1 d NS Seizures; weakness; short-term hearing deficit; cortical Jaffe et al. 1983

month old) atrophy

Limitations: Exposure data limited

BML Range: 16-43 μg/24 hr urine

Human/1 M ~2 hr NS Dizziness, weakness Lilis et al. 1985

Limitation: small sample size; limited exposure data

BML: 1,900 μg/L urine on first day

Human/1 F 2 mo NS Lethargy; irritability Foulds et al. 1987

(child) Limitations: small sample size; limited reporting of

symptoms; limited exposure data

BML: 214 μg/L in 24 hr urine

Human/1 M 2 wk NS Tremor; sleep disturbance; anxiety; cold hands and feet Karpathios et al.

(child) Limitation: small sample size; limited exposure data 1991

BML: 130 μg/24 hr urine

Human/17-26 <16 hr NS Fatigue, headaches, irritability, depression, anxiety, Bluhm et al. 1992a

M tremor, impaired performance on visual-motor tests

(p<0.05) reported in welders following accidental

exposure.

Limitation: Chronic exposure to other metals; exposure

data limited

BML: ~60 μg/L in blood at 20 d postexposure

Human/1 F 6 mo NS Peripheral neuropathy; erethism; dizziness; depression; Fagala and Wigg

(child) irritability 1992

Limitation: small sample size; exposure data limited

BML: 686 μg/24 hr urine

Human/2 F Several months 0.01-0.04 Numbness in fingers and toes; absence of deep tendon Taueg et al.

(children) several reflexes; visual field defects; weakness 1992

months after BML not reported

initial spill
3-13

Table 3-5

Neurotoxicity of Elemental Mercury in Humans: Epidemiological Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/27 3 mo-39 yr 0-1.67 161 electric meter repair workers were examined, and 22 Bidstrup et al. 1951

cases (sex NS) (occup) (est.) were found to be symptomatic; there were 5 index cases. 

Tremor; irritability; visual impairment were observed

Limitation: Concomitant exposure to other chemicals is

likely.

BML Range: 1,495-7,950 μg/24 hr urine

Human/3 M, 6 NS NS Neuropsychological tests showed irritability, tremor, memory Vroom and Greer

F exposed/10 (occup) loss, poor coordination, visual impairment; altered 1972

M, 30 F electrophysiology (p<0.05) in thermometer manufacturing

controls employees

Limitation: Exposure data limited

BML Range: 4-1,101 μg/24 hr urine

Human/43 >6 months NS Objectively assessed tremor and eye-hand coordination Roels et al. 1982

exposed/47 Mean: 5.3 yr tended to be higher in the exposed group, with a significant

controls (sex (occup) (p<0.05) difference on one test. There was a tendency toward

NS) a dose-response, but it was not statistically significant.

Exposed group worked in amalgam or chloralkali plants;

control workers were matched from the same plants, but

unexposed.

Limitation: Exposure data limited

BML: 29.2 μg/L in blood (range: 5.3-135); 95.5 μg/g

creatinine in urine (range 9.9-286)

Human/23 M NS (occup) NS Decreased nerve conduction velocity (p<0.05); visual Shapiro et al. 1982

exposed/22 M impairment (p<0.01); higher distress levels

control Of a sample of 298 dentists, the exposed group had "tissue"

mercury levels in the top 20%; the controls were agematched, with no detectable tissue mercury. Tissue mercury

in the head and wrist was measured using x-ray fluorescence.

Limitation: Exposure data limited

BML: >20 μg/g in tissue 

Human/12 3 mo-8 yr NS In a battery of objective tests, the following findings were Williamson et al.

exposed/12 (occup) significant: tremor (p<0.025); decreased verbal intelligence; 1982

controls (sex short- and long-term memory impairment (p<0.01); fatigue

NS) (p<0.01).

The exposed group worked with amalgam (n=4) or were

exposed to mercuric chloride, methoxyethyl mercuric

chloride, methoxyethylmercuric acetate (n=8). Controls were

matched by age, sex, education, ethnic background.

Limitation: Exposure data limited; small sample size

BML Range: <10-670 μg/L urine

Human/26 Avg: 15.3 yr 0.026 Objectively assessed tremor was significantly (p = 0.001) Fawer et al. 1983

exposed M/25 (occup) (TWA) elevated in the exposed group and correlated with exposure

control M personal duration. Exposed group worked in fluorescent tube factories

monitoring (n=7), chloralkali plants (n=12), or in acetaldehyde

production. The control subjects worked at the same

factories but had not been exposed to mercury. 

Mean BML: 8,280 μg/L in blood; 20 μg/g creatinine in urine

Human/36 M Avg: 16.9 yr 0.022– By comparison to age-matched controls, chloralkali workers Piikivi et al. 1984

exposed/36 (10-37 yr) 0.028 (est.) had memory impairment, decreased verbal intelligence

controls (occup) (p<0.01).

a

BML: >15μg/L in blood; >56 μg/L in urine
Table 3-5 (continued)

Neurotoxicity of Elemental Mercury in Humans: Epidemiological Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

3-14

Human/60 M Avg: 13.7 yr 0.025 (est.) In a psychological and psychomotor test battery, there were Piikivi and

exposed/60 M (5-28 yr) statistically significant differences in subjective tests Hanninen 1989

controls (occup) (memory disturbance, mood; p<0.01) and an objective test

(hand-eye coordination, p<0.001). Subjects were chlorinealkali workers and controls were age-matched.

BML: 10.4 μg/L avg. in blood; 17.9 μg/g creatinine avg. in

urine

Human/41 M 5–27 yr 0.025 (est.) Attenuation of power density spectrum of EEG in chloralkali Piikivi and Tolonen

exposed/41 M (occup) workers (p<0.01); controls were age-matched;slight increase 1989

controls in subjective symptoms of autonomic (cardiovascular)

dysfunction and a slight decrease in pulse rate variations

(cardiovascular reflex response).

BML: 67.8 μg/L in blood; 20.6 μg/g creatinine in urine

Human/54 M Avg: 7.7 yr NS Chloralkali and amalgam workers had impaired eye-hand Roels et al. 1989

exposed/48 (1-20 y coordination (p<0.001) and hand steadiness (p<0.02) in

controls (occup) objective tests, by comparison to unexposed matched controls

from the same plants.

Limitation: Exposure data limited

Geometric mean BML: 24 μg/L blood; 63 μg/g creatinine

Human/10 M, Avg: 5 yr 0.076 (avg) Neurological exam found difficulty with heel-to-toe gait Ehrenberg et al.

62 F exposed/9 (occup) 0.003-0.27 (p<0.05) in thermometer manufacturers; control population 1991

M, 60 F control (range) worked at a nearby electronics manufacturer.

Avg. BML: 73.2 μg/g creatinine in urine

Human/89 >1 yr 0.025 Increased tiredness, memory disturbance, based on interviews Langworth et al.

exposed/75 (occup) (p<0.001), but no effect on psychometric tests or tremor in 1992a

controls chloralkali workers.

BML: 11 μg/L in blood; 25.4 μg/g creatinine in urine

Human/60 M, 10 hr/d 0.014 Impaired performance on neurobehavioral tests in dentists Ngim et al. 1992

38 F 6 d/wk (TWA) (p<0.05); severity of effect correlated with exposure.

exposed/27 M, 0.7-24 yr Limitations: Concomitant exposure to physical and vibration

27 F controls (occup) load (affecting dexterity tests); confounding exposure to folk

medicine, BML: Mean 9.8 μg/L in blood; range: 0.63-57.3

μg/L in blood

Human/77 M >1 yr 0.059 In a study of ex-chloralkali workers (avg. 12.3 yr since last Ellingson et al. 1993

exposed/53 M Avg: 7.9 yr exposure) compared with age-matched controls, sensory

controls (occup) nerve conduction velocity and visual evoked response

correlated with mercury exposure (p<0.05)

BML: 3,190 μg/g creatinine in urine current, 106 μg/L in

urine during exposure

Human/117 M 389 min/d 1.5-3.3 Ex-mercury miners tested 18 years after the closure of the Kishi et al. 1993

exposed/76 duration NS mine had lower scores on objective neuropsychological tests

controls (occup) (motor coordination, reaction time, short-term memory) than

age- and education-matched controls (p<0.01). 76 of the

miners had a history of mercury poisoning, but subjective

symptoms had generally decreased since exposure

BML not reported
Table 3-5 (continued)

Neurotoxicity of Elemental Mercury in Humans: Epidemiological Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

3-15

Human/19 M, 2 yr 0.033 (avg) Increased fatigue and confusion; impaired performance on Liang et al. 1993

69 F Avg. 10.4 yr 0.008-0.085 neurobehavioral tests in fluorescent lamp factory workers

exposed/97 (occup) (range) (p<0.01)

controls Avg. BML: 25 μg/L in urine

 Estimate by extrapolating from urinary mercury levels (Roels et al. 1987). a

In animals, as in humans, adverse neurological effects are observed after exposure to elemental

mercury vapor. Effects observed in rabbits and mice after subchronic exposures included tremors, ataxia,

paralysis, failure to respond to light and decreased conditioned avoidance responding (Fukuda 1971;

Ganser and Kirschner 1985; Kishi et al. 1978). Pathologic changes (unspecified) were observed in the

brains of rabbits at 0.86 mg Hg/m (Ashe et al. 1953). 3

Table 3-6

Neurotoxicity of Elemental Mercury in Animals: Inhalation Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Rat/Albino/7 M 12-42 wk 0, 3 Tremor; decline in conditioned avoidance and conditioned Kishi et al. 1978

exposed/6 M 5 d/wk escape responses. First significant effect at 20 weeks

control 3 hr/d (p<0.05)

Limitation: Only one level tested

BML Range: 11.18-17.83 μg/g in cerebrum (wet weight)

Mouse/ 3.5 wk NS Ataxia; motor dysfunction Ganser and

C57BL6J/ 5 d/wk (saturated) Limitations: Poorly defined exposure conditions; limited Kirschner 1985

No. and sex NS 20–40 min/d data reporting on effects

BML not reported

Rabbit/774 1-12 wk 0.86 Mild to moderate pathological changes in brains Ashe et al. 1953

strain NS/ 5 d/wk Limitations: One exposure level; limited data reporting

31 (sex NS) 7 hr/d BML: Brain level 1.2 μg/g

Rabbit/strain 13 wk 0, 4 Tremor Fukuda 1971

NS/6 M 4 d/wk Limitation: No control

6 hr/d BML: 0.8–3.9 μg/g wet weight (brain)
3-16

3.1.3.3 Renal

The kidney is a sensitive target organ following inhalation exposure to elemental mercury. Acute

accidental exposure in private homes or as a result of industrial accidents resulted in symptoms ranging

from slight changes in urinary acid excretion to transient renal failure with proteinuria, nephrosis and

necrosis of the proximal convoluted tubules (Bluhm et al. 1992b; Jaffe et al. 1983; Rowens et al. 1991;

Tubbs et al. 1982). Proteinuria, proximal tubule damage and glomerulosclerotic changes were also

reported in a workers occupationally exposed for up to 2.5 years; in two cases the exposure levels were

measured at 0.02 to 0.45 mg/m (Friberg et al. 1953; Kazantzis et al. 1962). Comparisons of exposed 3

workers to unexposed controls found increased urinary N-acetyl--D-glucosaminidase in workers

exposed to 0.025 mg/m and increased incidence of proteinuria (Roels et al. 1982). 3

Table 3-7

Renal Toxicity of Elemental Mercury in Humans: Case Studies

Species/

No. per Exposure Dose

Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/2 M 8-9 mo 0.02-0.45 Proteinuria and nephrosis in electrochemical industry workers Friberg et al. 1953

(adult) (occup) Limitations: small sample size; concomitant exposure to

chlorine

BML: 160-900 μg/L in urine

Human/3 M 4 mo-2.5 yr NS Heavy albuminuria; transient renal failure; proximal tubule Kazantzis et al. 1962

(occup) damage; glomeruloscierotic changes

Limitations: small sample size; concomitant exposure to other

mercurials and other compounds; limited exposure data

BML Range: 1,100-1,440 μg/L in urine

Human/2 M/ NS (occup) NS Proteinuria; glomerulonephritis in chemical plant workers Tubbs et al. 1982

41 M controls Limitation: small sample size; concomitant exposure to other

metals; limited exposure data

BML Range: 174-548 μg/24 hr urine

Human/1 F 1 day NS Acute renal failure (proteinuria, glucosuria, granular casts) Jaffe et al. 1983

(8-month old) BML Range: 16 μg/24 hr urine

Human/ Once NS Necrosis of proximal tubule; increased serum urea nitrogen and Rowens et al.

2M, 2F creatinine in 2 subjects 1991

Limitations: small sample size; limited exposure data

BML Range: 94–220 μg/L Hg in urine

Human/11 M <16 hr NS Hyperchloremia, low normal bicarbonate in urine in welders Bluhm et al. 1992b

following accidental exposure

Limitations: No information on pre-exposure range

BML: ~60 μg/L in blood at 20 d postexposure
3-17

Table 3-8

Renal Toxicity of Elemental Mercury in Humans: Epidemiological Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/21 NS NS 0.01-0.05 Increased proteinuria in exposed pathology laboratory Stewart et al. 1977

(occup) workers compared to unexposed controls. Proteinuria cleared

when mercury exposure was limited.

BML: 35 μg/L in urine

Human/43 >6 months NS Proteinuria significantly elevated (p<0.05). Exposed group Roels et al. 1982

exposed/47 Mean: 5.3 yr worked in amalgam or chloralkali plants; control workers

controls (sex (occup) were matched from the same plants but unexposed.

NS) Limitations: Exposure data limited

BML: 29.2 μg/L in blood (range: 53-135); 95.5 μg/g

creatinine in urine (range 9.9-286)

Human/ 62 M 1-25 yr 0.046 (est.) Among exposed chloralkali plant or zinc-amalgam factory Lauwerys et al 1983

exposed, 60 M (avg 5.5 yr) workers, renal function parameters were not different from

controls (occup) unexposed controls. Circulating anti-laminin antibodies

found in eight exposed, 0 controls. No dose-effect

relationship between blood or urine levels and occurrence of

anti-laminin antibodies. Exposure level estimated using

Roels et al (1987) conversion factor.

BML: 16 μg/L (range 2.5-75.6) in blood; 56 μg/g creatinine

(range 3-272) in urine

Human/100 M 8 yr (avg) NS Small increase in prevalence of higher activities of NAG and Stonard et al. 1983

(occup) gamma-glutamyl transferase in chloralkali workers w/urinary

mercury excretion >100 μg/g creatinine. No renal function

changes in workers w/mean urine 67 μg/g creatinine

Limitation: Exposure data limited

BML: 67->100 μg/g creatinine in urine

Human/58 M 7.9 yr (avg.) 0.059 (est) Follow-up to Lauwerys et al. (1983). In contrast to the earlier Bernard et al. 1987

exposed study, there was no evidence of anti-laminin antibodies in

exposed workers

BML: 72 μg/g creatinine in urine

Human/41 M 1-20 yr 0, 0.025 Increased urinary N-acetyl--D-glucosaminidase in a group of Barregard et al.

exposed, 41 M (occup) chloralkali workers. Controls were age-matched. 1988

controls BML: 15.6 μg/L in blood

Human/ 60 M 13.7 5.5 yr NS No evidence of glomerular or tubular damage (effect on Pilkivi and

exposed, 60 M (occup) urinary albumin or N-acetyl--glucosaminidase activity) in Ruokonen 1989

controls chloralkali workers compared to controls. NOAEL of 25

+

mg/m based on Roels et al. (1987) conversion factor. 3

Limitation: Exposure data limited

BML: 14 μg/L in blood; 17 μg/L in urine

Only one study was found of kidney effects in animals from exposure to elemental mercury

vapor (Ashe et al. 1953). The observed effects supported the human data, with kidney effects ranging

from moderate unspecified pathological changes at shorter durations to necrosis and cellular

degeneration at longer durations. Limited quantitative data were reported.
3-18

Table 3-9

Renal Toxicity of Elemental Mercury in Animals: Inhalation Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Rabbit/strain 1-30 hr 28.8 Kidney pathology correlated with exposure duration, Ashe et al. 1953

NS/14 (sex ranging from moderate changes at 1 hour to widespread

NS) necrosis at 30 hours.

Limitations: No control group, limited data reporting

BML Range: 20-5,320 μg/L in blood

3.1.3.4 Respiratory

Respiratory toxicity in humans following exposure to elemental mercury vapors has been

characterized by pulmonary edema and congestion, coughing, interstitial pneumonitis, respiratory failure

and absence of air in lungs at time of histopathological examination (Bluhm et al. 1992a; Hallee 1969;

McFarland and Reigel 1978; Milne et al. 1970; Snodgrass et al. 1981; Taueg et al. 1992). One case of

occupational exposure to elemental mercury vapor occurred due to a faulty thermostat that heated to

450F and vaporized the mercury it contained. Signs included cough, chest pains, reduced vital capacity

and pneumonitis, which began within hours of the onset of exposure (McFarland and Reigel 1978). 

Accidental exposure to elemental mercury vapors in private homes has led to interstitial pneumonia,

dyspnea, lung disease and respiratory failure (Hallee 1969; Snodgrass et al. 1981; Taueg et al. 1992). In

each case, signs of toxicity persisted for days to months following acute exposure. No studies were

identified regarding respiratory effects in humans following intermediate or chronic exposures to

elemental mercury vapor.

Table 3-10

Respiratory Toxicity of Elemental Mercury in Humans: Case Studies

Species/

No. per Exposure Dose

Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/1 M, <12 hr NS Dyspnea; interstitial pneumonia; fibrosis; moderate restrictive Hallee 1969

1 F (adults) lung disease

Limitation: Case study

BML Range: 191-557 μg/24 hr urine

Human/4 M 2.5-5 hr 1.1-1.7 Cough; chest tightness occurred following an accidental Milne et al. 1970

(occup) (est.) exposure of electrochemical industry workers

Limitation: Case study

BML Range: 100-130 μg/L in urine 10-14 days postexposure
Table 3-10 (continued)

Respiratory Toxicity of Elemental Mercury in Humans: Case Studies

Species/

No. per Exposure Dose

Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

3-19

Human/6 M <8 hr 44.3 Pneumonitis; cough; chest pain McFarland and

(est.) Limitations: Case study; limited data reporting Reigel 1978

BML Range: 1,060-3,280 μg/24 hr urine

Human/2 M, 3 d NS Cough; dyspnea Snodgrass et al.

2 F (adults) Limitation: Case study 1981

BML: 82-5700 μg/24 hr urine

Human/1 M ~2 hr NS Reduced vital capacity and dynamic lung volumes, shortness of Lilis et al. 1985

breath

Limitation: Case study

BML: 1,900 μg/L urine on first day

Human/17 M <16 hr NS Congestion; dyspnea; lung infiltrates in up to 15/17 welders Bluhm et al. 1992a

interviewed following accidental exposure

Limitation: Limited data reporting of effects or exposure

BML: ~60 μg/L in blood 20 d postexposure

Human/2 M, ~24 hr NS Adult respiratory distress syndrome; respiratory failure Taueg et al. 1992

2 F (adults) BML Range: 4.6-219 μg/L in urine

Rats exposed to 27 mg Hg/m as elemental mercury vapor for one hour exhibited dyspnea, and 3

exposure for two hours resulted in death by asphyxiation (Livardjani et al. 1991). Histopathological

analyses revealed necrosis of the alveolar membrane, presence of hyaline membranes and evidence of

pulmonary edema. Acute-duration studies with rabbits revealed degeneration and necrosis of the lungs

(Ashe et al. 1953). Gage (1961) reported congestion and necrosis of the lungs following intermediateduration exposure to elemental mercury vapor at a concentration of 1 mg Hg/m .3
3-20

Table 3-11

Respiratory Toxicity of Elemental Mercury in Animals: Inhalation Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Rat/Wistar/ 7 wk 1 Congestion; necrosis of lung Gage 1961

6 F 100 hr/wk Limitation: Limited data reporting

5 d/wk BML: 10 μg/rat in lungs

Rat/Wistar/64 1 or 2 hr 0, 27 Death by asphyxiation; lung edema; hyaline membranes; Livardjani et al.

M/ duration necrosis of alveolar epithelium 1991

BML Range: 391-4,558 μg/L in blood

Rabbit/strain 1-30 hr 28.8 Pathology correlated with exposure duration and ranged Ashe et al. 1953

NS/14 (sex NS) from mild changes at 1 hour to marked cellular degeneration

and necrosis at 30 hours. In another study, Ashe reported

no respiratory damage in rats exposed to 0.1 mg/m3 for 72

weeks.

BML Range: 20-5,320 μg/L in blood

3.1.3.5 Cardiovascular

Signs of cardiovascular toxicity in humans after acute exposure to elemental mercury include

tachycardia, elevated blood pressure and heart palpitations (Bluhm et al. 1992a; Snodgrass et al. 1981;

Soni et al. 1992). Intermediate-duration exposure to elemental mercury vapors produced similar effects

(i.e., tachycardia and elevated blood pressure) (Fagala and Wigg 1992; Foulds et al. 1987). Barregard et

al. (1990) performed a study on chloralkali workers and showed that they had an increased risk of

ischemic heart disease and cerebrovascular disease. These workers, however, were exposed to other

chemicals and to magnetic fields which may have affected the results. Piikivi (1989) demonstrated a

positive correlation between heart palpitations and urinary mercury concentrations in workers from a

chloralkali plant. It is unclear from the available scientific literature, however, whether the effects on

cardiovascular function (e.g., tachycardia, elevated blood pressure) are due to direct cardiac toxicity or to

indirect toxicity (e.g., due to effects on neural control of cardiac function) of elemental mercury.
3-21

Table 3-12

Cardiovascular Toxicity of Elemental Mercury in Humans: Case Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/2 M, 2 3 d NS Elevated blood pressure; tachycardia Snodgrass et al.

F (adults) Limitations: Case study; limited exposure data 1981

BML Range: 82-5,700 μg/24 hr urine

Human/1 F 2 mo NS Elevated blood pressure; tachycardia Foulds et al. 1987

(child) Limitations: Case study; limited exposure data

BML Range: 214-296 μg/L in 24 hr urine

Human/17 M <16 hr NS Palpitations in 5/17 welders interviewed following accidental Bluhm et al. 1992a

exposure

Limitation: Exposure data limited

BML: ~60 μg/L in blood 20 d postexposure

Human/1 F 6 mo NS Elevated blood pressure; tachycardia Fagala and Wigg

(child) Limitations: Case study; limited exposure data 1992

BML: 686 μg/24 hr urine

Human/1 M (3- <6 hr NS Tachycardia Soni et al. 1992

yr old) Limitations: Case study; limited exposure data

BML not reported

Table 3-13

Cardiovascular Toxicity of Elemental Mercury in Humans: Epidemiological Studies

Species/

No. per Exposure Dose

Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/41 M 16 yr (avg) 0.03 Palpitations in chloralkali workers (p<0.05); no significant effect Piikivi 1989

exposed/41 5-27 yr (est.) on cardiovascular reflex responses compared to matched controls.

M controls (occup) BML Range: 3.5-52.5 μg/L in urine; avg 19.3 μg/L in urine

Human/ 10 yr (avg) Avg Increased mortality due to ischemic heart and cerebrovascular Barregard et al.

26 M (occup) samples: disease in chloralkali workers, compared to matched controls. 1990

0.025- Limitation: Possible confounding due to shift work

0.050 BML: Decrease from 200 μg/L in urine in 1950's to <50 μg/L in

1990

Few animal studies were located regarding cardiovascular effects after exposure to elemental

mercury vapor. Studies in rabbits report unspecified cellular degeneration and necrosis of the

cardiovascular system following both acute and intermediate exposure (Ashe et al. 1953). Ashe et al.

(1953), however, concluded that the concentration of mercury is a better indicator of cardiovascular

toxicity than the duration of exposure, especially at lower exposure levels.
3-22

Table 3-14

Cardiovascular Toxicity of Elemental Mercury in Animals: Inhalation Exposure

Species/

Strain/

No. per

Sex per Exposure Dose

Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Rabbit/strain 1-30 hr 28.8 Pathology correlated with exposure duration and ranged from Ashe et al. 1953

NS/14 (sex mild changes in the heart at 1 hour to marked cellular

NS) degeneration and necrosis at 12 hours.

Limitations: No controls; limited data reporting;

only one dose level tested

BML Range: 20-5,320 μg/L in blood

Rabbit/strain 1-11 wk 6 Mild to moderate pathological changes of the heart. Pathological Ashe et al. 1953

NS/16 (sex 5 d/wk changes observed in subchronic studies were correlated with the

NS) 7 hr/d exposure concentration but not exposure duration.

Limitations: No controls; limited data reporting;

only one dose level tested

BML Range: 70-3,000 μg/L in blood

Rabbit/strain 12 wk 0.86 Most animals had mild heart pathology, but 2 animals each at 6 Ashe et al. 1953

NS/31 (sex 5 d/wk and 7 weeks had marked cellular degeneration and necrosis, with

NS) 7 hr/d focal fibrosis.

Limitations: Limited data reporting; only one dose level

BML Range: 50-620 μg/L blood

3.1.3.6 Gastrointestinal

Gastrointestinal effects have been reported by persons exposed to elemental mercury vapor. The

most common sign of mercury poisoning is stomatitis (inflammation of the oral mucosa), which is

usually reported following acute, high concentration exposure to elemental mercury vapors (Bluhm et al.

1992a; Snodgrass et al. 1981). Sexton et al. (1978), however, reported signs of bleeding gingiva in 12

people exposed to mercury vapors for two months after metallic mercury was spilled in two homes, and

Schwartz et al. (1992) reported bleeding gums in a child exposed to mercury vapors for two to four

weeks. In addition, Vroom and Greer (1972) documented mercury intoxication in nine workers at a

thermometer manufacturing plant; the workers complained of sore gums and lesions on the oral mucosa

after long-term exposure. Other commonly reported gastrointestinal effects include nausea, vomiting,

diarrhea and abdominal cramps (Bluhm et al. 1992a; Campbell 1948; Lilis et al. 1985; Sexton et al. 1978;

Snodgrass et al. 1981; Vroom and Greer 1972).
3-23

Table 3-15

Gastrointestinal Toxicity of Elemental Mercury in Humans: Case Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/1 F 5 hr NS Difficulty swallowing; abdominal pain; necrosis of stomach Campbell 1948

(4-month old) mucosa and duodenum

Limitations: Case study; exposure data limited

BML not reported

Human/3 M, NS NS Sore gums; diarrhea in thermometer manufacturing employees Vroom and Greer

6 F (occup) Limitation: Exposure data limited 1972

BML Range: 4-1,101 μg/24 hr urine

Human/5 M, 51-176 d 0.1-1.0 Nausea, vomiting, abdominal pain, anorexia, diarrhea, bleeding Sexton et al. 1978

6 F (adults and gingiva more common than in controls

children)/12 Limitations: Small sample size; no statistical analysis

controls (sex BML avg: 3.7 μg/L in urine;

NS) BML Range: 183-620 μg/L in blood

Human/2 M, 3 days NS Nausea; vomiting; swelling of gums Snodgrass et al.

2 F (adults) Limitation: Case study 1981

BML Range: 13-5,700 μg/24 hr urine

Human/1 M 2 hr NS Nausea; vomiting Lilis et al. 1985

Limitations: Case study; limited reporting of symptoms;

exposure data limited

BML Range: 900-1,900 μg/L in urine (over 3 days)

Human/17 M <16 hr NS Diarrhea; cramps in up to 11/17 welders accidentally exposed Bluhm et al. 1992a

Limitations: Case study; exposure data limited

BML not reported

Very little information is available concerning gastrointestinal toxicity after exposure to

elemental mercury vapors. Ashe et al. (1953) exposed rabbits to mercury vapors for 1–30 hours at a

concentration of 28.8 mg Hg/m and found unspecified cellular degeneration and necrosis. When rabbits 3

were exposed to 6 mg Hg/m for 1–11 weeks, changes in the colon were seen during histopathological 3

analysis (Ashe et al. 1953).
3-24

Table 3-16

Gastrointestinal Toxicity of Elemental Mercury in Animals: Inhalation Exposure

Species/

Strain/

No. per

Sex per Exposure Dose

Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Rabbit/strain 1-30 hr 28.8 Earliest effect on colon (mild pathological changes) occurred at 2 Ashe et al. 1953

NS/14 (sex hr; marked cellular degeneration and necrosis observed at 30 hr

NS) Limitations: No controls; limited data reporting

BML Range: 20-5,320 μg/L in blood

Rabbit/strain 1-11 wk 6.0 No effects or mild histopathological changes in colon Ashe et al. 1953

NS/16 (sex 5 d/wk Limitations: No controls; limited data reporting

NS) 7 hr/d BML Range: 70-3,600 μg/L blood

3.1.3.7 Hepatic

Biochemical changes in hepatic enzymes were noted in a child who was exposed for

approximately one day to an unspecified concentration of elemental mercury vapors (Jaffe et al. 1983). 

Serum glutamic-pyruvic transaminase (SGPT) and bilirubin levels were elevated, and synthesis of

hepatic coagulation factors was reduced. No human studies were identified regarding the hepatic toxicity

of mercury following intermediate or chronic exposures to elemental mercury vapors.

Table 3-17

Hepatic Toxicity of Elemental Mercury in Humans: Case Study

Species/

No. per Exposure Dose

Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/1 F 1 d NS Elevated serum alanine amino-transferase and bilirubin Jaffe et al. 1983

(8-month old) Limitations: Case study; exposure data limited

BML: 16 μg/24 hr urine

Ashe et al. (1953) performed histopathological analyses on rabbits after exposing them for one to

30 hours or for one to 11 weeks to elemental mercury vapors. The analyses revealed necrosis and

cellular degeneration of the liver. No other animal studies were identified regarding the hepatic toxicity

of mercury vapors following inhalation exposure.
3-25

Table 3-18

Hepatic Toxicity of Elemental Mercury in Animals: Inhalation Exposure

Species/

Strain/

No. per

Sex per Exposure Dose

Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Rabbit/strain 1-30 hr 28.8 Pathology correlated with exposure duration. Moderate changes Ashe et al. 1953

NS/14 (sex first occurred at 2 hr and widespread necrosis at 30 hr

NS) Limitations: No controls; limited data reporting

BML Range: 20-5,320 μg/L in blood

Rabbit/strain 1-11 wk 6.0 Pathology was somewhat correlated with exposure duration and Ashe et al. 1953

NS/16 (sex 5 d/wk ranged from mild to marked cellular degeneration with necrosis.

NS) 7 hr/d Limitations: No controls; limited data reporting

BML Range: 70-3,600 μg/L blood

3.1.3.8 Hematological

After acute-duration exposure to high concentrations of elemental mercury vapor, onset of "metal

fume fever" may occur; this syndrome is characterized by leukocytosis with fever, chills and fatigue

(Campbell 1948; Haddad and Stenberg 1963; Jaffe et al. 1983). Intermediate-duration exposure to

mercury vapors led to an elevated white blood cell count in a 12-year-old female after exposure for six

months (Fagala and Wigg 1992). Volunteers with dental amalgams had significantly decreased

hemoglobin and hematocrit compared to controls without dental amalgams (Siblerud 1990).

Table 3-19

Hematological Toxicity of Elemental Mercury in Humans: Case Studies

Species/

No. per Exposure Dose

Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/1 F 6 mo NS Elevated white cell count Fagala and Wigg

(child) Limitations: Case study; exposure data limited; skin lesions 1992

could have led to elevated count

BML: 686 μg/24 hr urine

Human/1 M 2-4 wk NS Thrombocytopenia Schwartz et al. 1992

(3.5 yr old) Limitation: Exposure data limited

BML: 151 μg/L in blood
3-26

Table 3-20

Hematological Toxicity of Elemental Mercury in Humans: Epidemiological Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/47 (sex NS <0.1 Decreased -aminolevulinic acid dehydratase and Wada et al. 1969

NS) (occup) cholinesterase activity in erythrocytes, effects were

significantly (p<0.01) correlated to urinary mercury

BML Range: 2-472 μg/g of creatinine in urine.

Human/41 M NS Range: Increased 2-macroglobulin and ceruloplasmin in mercury Bencko et al. 1990

exposed/55 (occup) 0.106- plant workers compared to unexposed controls (p<0.001)

controls 0.783 BML Range: 29-545 μg/L in urine 

Human/20 M, NS NS Subjects with amalgams had decreased (p<0.02) mean Siblerud 1990

30 F hemoglobin (14.66±1.09 g/dL in subjects vs. 14.88±1.14 g/dL

exposed/21 M, in controls) and mean hematocrit (43.15±3.66% in subjects vs.

30 F control 43.91±3.61% in controls). These reductions were significantly

(p<0.01) correlated with increasing urine mercury in the

subjects with amalgam.

Limitations: Subjects identified through newspaper ads may

have introduced self-selection bias; only mean data reported.

BML avg: 3.7 μg/L in urine

No animal studies were identified regarding the hematological toxicity of mercury vapors

following inhalation exposure.

3.1.3.9 Immunological

The available evidence suggests that the immune reaction to elemental mercury exposure is

idiosyncratic, with either increases or decreases in immune activity depending on genetic predisposition. 

Although there is evidence for an overall suppression of the humoral immune response among exposed

workers (Moszczynski et al. 1990), this effect has not been consistently observed (Bencko et al. 1990;

Langworth et al. 1992b). The failure to observe consistent decreases in antibody content of the serum

may be due to small numbers of workers in each group who develop an autoimmune reaction upon

exposure to mercury. For example, small numbers of workers exposed to elemental mercury vapors have

had elevated levels of antiglomerular basement membrane and anti-DNA antibodies (Cardenas et al.

1993; Langworth et al. 1992b) or granular deposition of IgG and complement C3 in the renal glomeruli

(Tubbs et al. 1982).
3-27

Table 3-21

Immunotoxicity of Elemental Mercury in Humans: Case Study

Species/

No. per Exposure Dose

Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/2 M (occup) <0.1 BML Range: 174-548 μg/24 hr urine Tubbs et al. 1982

NS Limitation: Case study

Deposition of IgG and C3 in glomeruli of chemical plant workers

Table 3-22

Immunotoxicity of Elemental Mercury in Humans: Epidemiological Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/41 M NS Range: Decreased IgG; increased IgA and IgM in mercury plant Bencko et al. 1990

exposed/55 (occup) 0.106- workers (p<0.05)

controls 0.783 BML Range: 29-545 μg/L in urine

Human/50 1.5-25 yr NS Abnormally high anti-DNA antibody titre (p<0.01) Cardenas et al. 1993

exposed/50 Avg: 11 yr Mean BML: 31.9 μg/L in urine

controls (occup)

An autoimmune response to mercury has been produced in a susceptible strain of rats (Brown

Norway) exposed to mercury vapor (Hua et al. 1993). In these rats, increased levels of serum IgE and

antilaminin autoantibodies, deposition of IgG deposits in the renal glomeruli and proteinuria were

observed.

Table 3-23

Immunotoxicity of Elemental Mercury in Animals: Inhalation Exposure

Species/

Strain/

No. per

Sex per Exposure Dose

Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Rat/BN/3-4 5 wk 0, 1 Increased serum IgE; anti-laminin autoantibody titre, IgG Hua et al. 1993

M, 3-4 F 6 or 24 hr/d deposits along glomerular capillary walls (p<0.001)

Mean BML: 90.3 μg/L in blood
3-28

3.1.3.10 Dermal

Exposure to elemental mercury vapors for acute or intermediate durations may elicit a response

known as acrodynia or "pink disease", which is characterized by peeling palms of hands and soles of feet,

excessive perspiration, itching, rash, joint pain and weakness, elevated blood pressure and tachycardia

(Fagala and Wigg 1992; Karpathios et al 1991; Schwartz et al 1992). Children seem to be the most

susceptible to acrodynia, although adults may be affected to a lesser degree (Warkany and Hubbard

1953). One man experienced a rash and stomatitis after inhalation exposure to mercury when repairing a

cell in a chloralkali plant (Bluhm et al. 1992a); however, dermal exposure may have also occurred.

Table 3-24

Dermal Toxicity of Elemental Mercury in Humans: Case Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/1 M 2 wk NS Red palms and soles; perspiration; rash Karpathios et al.

(child) Limitations: Case study; concomitant dermal exposure 1991

possible; exposure data limited

BML: 130 μg/24 hr urine

Human/17 M <16 hr NS Conjunctivitis; dermatitis in 8/17 welders exposed in an Bluhm et al. 1992a

accident

Limitation: Exposure data limited

BML not reported

Human/1 F 6 mo NS Peeling skin on palms and soles Fagala and Wigg

(child) Limitations: Case study; exposure data limited 1992

BML: 686 μg/24 hr urine

Human/1 M 2-4 wk NS Maculopapular whole body rash Schwartz et al. 1992

(3-yr old) Limitations: Case study; exposure data limited

BML: 151 μg/L in blood

No animal studies were identified regarding the dermal toxicity of mercury vapors following

inhalation exposure.

3.1.3.11 Developmental

Although few reports have addressed the effects of maternal exposure to elemental mercury

vapor on the developing fetus, the available information suggests that maternal exposure to sufficiently

high concentrations of elemental mercury vapor may adversely affect the developing fetus. A study of

the pregnancies of Polish dental professionals showed a high frequency of malformations of a

nonspecified nature (Sikorski et al. 1987). In contrast, a study of Swedish dental professionals found no

increases in malformations, abortions, or stillbirths (Ericson and Kallen 1989). An increase in low birth

weight infants was noted in the offspring of female dental nurses (Ericson and Kallen 1989); however, in

this same study similar effects were not observed for either dentists or dental technicians, and

socioeconomic factors may have contributed to the effects observed. It is unknown to what extent

discrepancies in the results of the above studies are attributable to differences in mercury exposure levels

(only the study by Sikorski et al. (1987) attempted to assess exposure levels) or to other confounders.
3-29

Table 3-25

Developmental Toxicity of Elemental Mercury in Humans: Case Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/1 F 8 mo NS Infant death at birth; fetal hepatomegaly; spontaneous Derobert et al. 1950

(occup) abortion in 2 successive pregnancies of a thermometermanufacturing worker. Maternal toxicity included tremors,

motor incoordination, hyperreflexivity, stomatitis.

Limitations: Case study; exposure data limited; maternal

toxicity also occurred.

BML not reported

Human/1 F 2 yr NS Delivery of viable infant at term to thermometer factory Melkonian and

(occup) worker with mild peripheral neuropathy attributed to mercury. Baker 1988

Maternal toxicity included slight decrease in sensory reflexes.

Limitations: Case study; exposure data limited; no

neurological assessment of infant; slight maternal toxicity also

reported.

BML: Mother: 875 μg/L in urine; Offspring: 2.5 μg/L in

urine

Human/1 F 17 wk 0.02-0.06 Delivery of normal child who met all developmental Thorp et al. 1992

milestones. No maternal toxicity reported.

Limitation: Case study; no psychodevelopmental testing

BML: Mother: 230 μg/L in 24 hr urine at 17 wk, then

declined; Offspring: 3,000 μg/g in hair

The few animal studies that were identified indicate that inhalation of elemental mercury vapor

may be toxic to the developing animal. In an abstract, Steffek et al. (1987) reported decreased fetal

weight in offspring of rats exposed to elemental mercury vapor during gestation. Increased fetal and

postnatal deaths were also reported by Baranski and Szymczyk (1973) among rats exposed to elemental

mercury vapor for three weeks prior to mating and then again on gestation days 1–20, and increased

resorptions in rats exposed on gestation days 10–15 or 1–20.

Pregnant Sprague-Dawley rats (12/group) were exposed on gestation days 11–14 and 17–20 to

elemental mercury vapors (1.8 mg/m ) for one or three hours/day (Danielsson et al. 1993). Litters were 3

culled to 4 males and 4 females. Behavioral testing was done on one male and one female adult from

each litter; the authors state that for behavioral testing 8 were tested for each group. There was no

difference between controls and treatment groups for maternal weight gain. There was no obvious

mercury toxicity in the dams. Offspring exposed in utero were no different from controls in the following

measures: body weight; clinical signs; pinna unfolding; surface righting reflex development; tooth

eruption; and results of a negative geotaxis test at days 7, 8 or 9 post partum. At 3 months of age,

exposed male but not female rats showed significant decrements in four measures of spontaneous motor

activity: locomotion, rearing, rearing time and total activity. By 14 months, the high-dose animals

showed hyperactivity in the same test. Females were not evaluated in other adult behavioral tests. A test

for habituation to novel environment at 7 months of age showed significant differences between controls

and treated males on four measures. At 4 months, mercury-treated males had significantly higher latency

in a maze learning test; at 15 months, there was no difference between controls and treated rats in a

circular swim maze test.
3-30

Table 3-26

Developmental Toxicity of Elemental Mercury in Humans: Epidemiological Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/ 349 F NS NS Rates of pregnancy and labor complications were high among Mishinova et al.

exposed, 215 F (occup) women exposed to elemental mercury. Insufficient detail 1980

controls provided to evaluate dose-response relationship.

Limitation: Lack of exposure or effect data

BML not reported

Human/57 F 0.5-27 yr NS In a study of 57 dental professionals (117 pregnancies), Sikorski et al. 1987

(occup) reproductive failure (spontaneous abortion, stillbirth or

congenital malformation--not described further) was higher

than among unexposed controls, and the effect correlated

with exposure level (p=0.004). No maternal toxicity

reported.

Limitations: Small study group; control group not described;

exposure data limited

BML avg: 0.527 μg/g in scalp hair

Human/8157 F NS NS Study of infants born to dental workers, compared with the Ericson and Kallen

exposed (occup) general population. Based on medical registry, no increase in 1989

malformations, abortions, or stillbirths. Increased incidence

of low birth weight infants among offspring of dental

assistants (risk ratio 1.2, 95% C.I. 1.0-1.3), but the risk ratio

was decreased for dentists, suggesting a socioeconomic

effect. Case-control study of infants with neural tube defects

found none born to dentists, but the expected number was

only 0.5. No maternal toxicity reported.

Limitation: Exposure data limited

BML not reported.

Early postnatal exposure (during a period of rapid brain growth) resulted in subtle behavioral

changes when the rats were tested as young adults (Fredriksson et al. 1992). Eight litters/group, culled to

8 individuals, were exposed to 0.05 mg/m for either 1 or 4 hr/day. Exposure was on days 11–17 of age. 3

There were no signs of overt toxicity or changes in body weight. Spontaneous motor activity was

evaluated at 2 and 4 months. The high-dose group showed increased rearing at the early test, but the

repeat test indicated hypoactivity. The low-dose group was no different from controls at two months; at

four months this group showed increased total activity and decreased rearing. In the spatial learning test

administered at 6 months, low- dose rats had increased time to complete the task. High-dose animals

were observed to have increases in time to complete the task and in numbers of errors. No information

was given on the number of males and females tested or on any differences in behavior dependent on

gender.
3-31

Table 3-27

Developmental Toxicity of Elemental Mercury in Animals: Inhalation Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Rat/Strain Group I: 6 hr/d, 0, 2.5 Group I: Decreased number of live pups (p<0.05); Baranski and

NS/23-24 F 6-8 weeks decreased relative kidney (p<0.01) and liver weights Szymczyk 1973

before (p<0.05) and increased ovaries (p<0.05) in 2-month-old

fertilization: pups. Group II: Mean number of live fetuses lower than

Group II: 3 wk Limitations: Wide range in actual mercury concentration

before mating (0.5-4.8 mg/m ); only one level tested; maternal toxicity

and Gd 7-20 BML not reported

in controls

3

Rat/Sprague- 6 or 20d 0, 0.1, 0.5, Increased resorptions (LOAEL = 0.5 for Gd 10-15 and Steffek et al. 1987

Dawley/NS F 24 hr/d 1.0 1.0 for Gd 1-20); decreased maternal and fetal weights in

Gd 10-15 or group exposed to 1.0 on Gd 1-20.

Gd 1-20 Limitations: Reported only as an abstract; limited study

details; maternal toxicity

BML not reported

Rat/Sprague 1 or 3 hr/day on 0, 1.8 mg/m Hypoactivity at 3 months; hyperactivity at 14 months; Danielsson et al.

Dawley/ 4M, 4F Gd 11-14 plus decrement in habituation to novel environment at 7 1993

17-20 months; retarded learning in radial arm maze at 4 months

3

but no difference from controls in circular swim maze at

15 months. BML ranges for control through high dose

group (mg Hg/kg in organs): 0.001-0.012 (brain); 0.004-

0.112 (liver); 0.002-0.068 (kidney).

Limitations: Limited testing of female offspring; no

evaluation of differences between males and females;

small numbers of rats/group.

Rat/Sprague- 7 d 0, 0.05 Impaired spatial learning at 6 months (p0.01); increased Fredriksson et al.

Dawley/8 F 1 or 4 hr/d locomotor activity in objective test (p0.01) 1992

on post-partum BML Range: 0.017-0.063 μg/g in brain

days 11-17 Limitations: No information on gender-specific

behavioral effects; small number of animals/group.

Squirrel 4-7 hr/day, 5 0.5 or 1.0 Instability in lever-press durations and steady-state Newland et al. 1996.

monkey/10M, 1F d/week, during (1304 to performance under concurrent schedules; aberrant

gestation 4305 g transitions in treated animals. Five male monkeys born at

total) same time served as controls; there were 5 treated M and

one treated F. Maternal BML ranges 0.025 to 0.18 g/g.

3.1.3.12 Reproductive

Most studies that have examined the effects of occupational exposure to elemental mercury

vapor on reproductive function have failed to find evidence of adverse effects (Alcser et al. 1989;

Brodsky et al. 1985; Erfurth et al. 1990; Ericson and Kallen 1988; Heidam 1984; Lauwerys et al. 1985;

McGregor and Mason 1991). A few studies have shown at least suggestive evidence that elemental

mercury exposure may adversely affect reproductive function. In females exposed occupationally to

metallic mercury vapor, a correlation was observed between scalp hair mercury and reproductive failure
3-32

or menstrual abnormalities (Sikorski et al. 1987). An increased incidence of pregnancy complications

such as toxicosis or prolonged or hemorrhagic parturition was observed in exposed females when

compared to unexposed controls (Mishonova et al. 1980). A slightly increased incidence of menstrual

disorders in exposed females was reported by DeRosis et al. (1985); however, the statistical significance

of this finding was not presented. No evidence for an effect on fertility was observed in exposed males,

but one study of wives of exposed workers found an increased rate of spontaneous abortions (Cordier et

al. 1991). It is possible that the wives were exposed to mercury as the result of handling contaminated

clothing. None of the above studies presented information on exposure levels, and few presented

biomonitoring data. Thus, it is difficult to compare findings in the various studies.

Table 3-28

Reproductive Toxicity of Elemental Mercury in Humans: Epidemiological Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/728 F NS NS No increase in rate of spontaneous abortions in gardeners, Heidam 1984

exposed/1034 F (occup) dental assistants, painters, and factory workers.

controls BML not reported

Human/29,514 NS NS No correlation between mercury exposure (low and high) and Brodsky et al. 1985

M, 30,272 F (occup) rate of spontaneous abortions in dentists, dental assistants, or

their wives.

BML not reported

Human/153 F <5-17 yr <0.01 Slightly increased prevalence of menstrual disorders in DeRosis et al. 1985

exposed/193 F (occup) TWA at mercury lamp manufacturers, compared with workers subject

controls study; to similar stresses but not exposed to mercury.

>0.05 for 4 Limitations: Subjective measures; no statistical analysis

yr BML not reported

Human/103 M Avg: 5.9 yr NS No effect of paternal exposure on fertility of chloralkali, Lauwerys et al. 1985

exposed/101 M (1–25 yr) amalgam or electrical equipment workers, compared to

controls (occup) controls with similar workloads.

Avg BML: 52.4 μg/g creatinine in urine (range 5.1-272)

Human/57 F 0.5-27 yr NS In a study of 57 dental professionals (117 pregnancies), Sikorski et al. 1987

exposed (occup) reproductive failure (spontaneous abortion, stillbirth or

congenital malformation) was higher than among unexposed

controls, and the effect was correlated with exposure level

(extrapolated from hair Hg levels) (p=0.004). Irregular,

painful, or hemorrhagic menses was correlated with exposure

duration (p=0.005).

Limitations: Small study size; control group not described;

exposure data limited

Avg BML: 0.527 μg/g in scalp hair

Human/8157 F NS NS Based on medical registry, there was no increase in Ericson and Kallen

(occup) spontaneous abortions or stillbirths in pregnancies of dental 1988

professionals, compared to the general population.

BML not reported

Human/247 M 4 mo–8 yr NS No association between paternal exposure and rate of Alcser et al. 1989

exposed/255 M (occup) miscarriages in Department of Energy plant workers.

controls Limitation: Potential recall bias

BML: reported only as value integrated over time
Table 3-28 (continued)

Reproductive Toxicity of Elemental Mercury in Humans: Epidemiological Studies

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

3-33

Human/20 M 2–18 yr NS No correlation between blood or urinary mercury, and male Erfuth et al. 1990

exposed/21 M (occup) gonadotropic hormones of chloralkali workers, other

controls industrially-exposed workers, or dentists and matched

controls.

BML: Avg. 46 μg/g creatinine in urine (workers); 2.3 μg/g

creatinine (dentists)

Human/152 F NS NS Increased spontaneous abortions in women whose husbands Cordier et al. 1991

exposed/374 F (occup) were exposed to mercury vapors in chloralkali plants (rate

controls doubled above 50 μg/L in urine; 95% C.I. = 0.99-5.23)

Limitation: Exposure data limited

Avg BML: 61.9 μg/L in urine (range 26.9–75.9 μg/L)

Human/ 2–20 yr NS No correlation between blood or urinary Hg and male McGregor and

40 M exposed/ (occup) gonadotropic hormones in workers from different industries Mason 1991

63 M controls (not specified)

Avg BML: 103 μg/g creatinine in urine

In rats exposed to elemental mercury vapor, prolongation of estrous cycles was observed both

when compared to either unexposed controls or preexposure rates of cycling (Baranski and Szymczyk

1973).

Table 3-29

Reproductive Toxicity of Elemental Mercury in Animals: Inhalation Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Rat/Strain 3 wk 2.5 Longer estrous cycles, but the effect was not statistically Baranski and

NS/24 F 5 d/wk significant Szymczyk 1973

6 hr/d and BML not reported

Gd 7-20

3.1.3.13 Genotoxicity

Cytogenetic monitoring studies in populations exposed occupationally to elemental mercury

vapor provide conflicting evidence for a clastogenic effect of elemental mercury. Early studies reported

increased frequencies of chromosomal aberrations among exposed workers (Popescu et al. 1979;

Verschaeve et al. 1976). These studies, however, were not well-controlled, and the results could not be

reproduced in later studies (Mabille et al. 1984; Verschaeve et al. 1979). Popescu et al. (1979) compared

two groups of men exposed to elemental mercury vapor (Group I, n=4; Group II, n=18) with an
3-34

unexposed group of ten individuals and found a statistically significant increase in incidence of

chromosome aberrations in the exposed groups. Verschaeve et al. (1976) found an increase in

aneuploidy in lymphocytes of 28 subjects exposed to low concentrations of mercury vapor (by

comparison to seven controls), but these results were not repeated in later studies (Verschaeve et al.

1979). Mabille et al. (1984) did not find increases in structural chromosomal aberrations of lymphocytes

of exposed workers.

More recently, Barregard et al. (1991) demonstrated a correlation between cumulative mercury

exposure and induction of micronuclei among a group of chloralkali workers, suggesting a clastogenic

effect. This study did not show significant differences in frequency or size of micronuclei between the

exposed group to unexposed controls who were matched for age and smoking habits. Neither did they

find a correlation between the induction of micronuclei and current mercury exposure as measured by

blood or urine mercury levels. A correlation, however, was observed between cumulative exposure to

mercury and micronuclei induction in T-lymphocytes in exposed workers suggesting a genotoxic effect.

Table 3-30

Genotoxicity of Elemental Mercury in Humans

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/ NS NS Aneuploidy was significantly (p<0.001) increased in Verschaeve et al.

8 M, 6 F (exposed)/ (occup) subjects exposed due to an unstated occupation or as a 1976

3 M, 4 F (control) result of an accident at a university. Structural aberrations

were not increased.

Limitations: Small study size; smoking status not reported

BML Range: 1-114 μg/L in urine

Human/4 M 9.25 yr (avg) Range: Chromosome breaks (excluding gaps) were significantly Popescu et al. 1979

exposed/10 controls (occup) 0.15-0.44 (p<0.001) increased in whole blood cultures taken from

(sex NS) chemical plant workers. There was no effect on numerical

chromosome aberrations.

Limitations: Small sample size; smoking status not

reported.

BML Range: 142-386 μg/L in urine at study

Human/ 1-11 yr <0.05 at The incidence of structural and numerical chromosome Verschaeve et al.

28 exposed/ (occup) time of aberrations in exposed chloralkali plant workers did not 1979

20 controls (sex NS) study differ from controls. Eight of the controls were unexposed

workers at the same plant and 12 were taken from the

general population.

Limitations: Small sample size; there were 12 smokers in

the exposed group, 4/8 among the internal controls, and an

unknown number of smokers in the external controls

Avg BML: 35.4 μg/L in urine

Human/ 4 yr (avg) NS No increase in structural chromosome aberrations in zinc Mabille et al. 1984

22 exposed/ (0.3-15.3 yr) amalgam or chloralkali workers, compared to unexposed

25 controls (sex NS) (occup) control workers at the same plant.

Limitations: Small sample size; there were 15 smokers in

the exposed group and 12 among the controls; limited

exposure data

Avg BML: 30.6 μg/L in blood; range: 7.5-105 μg/L
Table 3-30 (continued)

Genotoxicity of Elemental Mercury in Humans

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

3-35

Human/26 M 10 yr (avg) Avg The frequency of micronuclei in lymphocytes was Barregard et al.

 (min. 1 yr) samples: correlated with cumulative exposure in Swedish chloralkali 1991

 (occup) 0.025- workers (p = 0.0035). There was no significant difference

0.050 between the frequency in the exposed and control

populations.

Controls were matched by age; exposed and control groups

each had 14 smokers

Limitation: Small study size

Avg BML: 9.6 μg/L in blood

No studies were identified that examined the genotoxicity of elemental mercury in animals

following inhalation exposure. Likewise no studies of genotoxic effects of mercury exposure in vitro

were recovered.

3.2 Inorganic Mercury

Inorganic mercury occurs in numerous forms/compounds; the most common include mercuric

chloride (HgCl ), mercurous chloride (Hg Cl ), mercuric oxide (HgO). The tables in this section include 2 22

a notation in the dose column indicating the specific form of inorganic mercury involved in that study. 

Oral doses, shown in mg/kg-day, have been converted to mg Hg/kg-day using the method shown in 

Appendix A.

3.2.1 Critical Noncancer Data

This section describes studies evaluated by U.S. EPA for use in assessing general systemic health

risks. Chapter 6 describes the derivation of an oral Reference Dose (RfD) for inorganic mercury based

on several studies wherein kidney diseases consequent to immunological effects were observed. For

completeness, some of these studies are also presented in tabular form in succeeding sections.

3.2.1.1 Human Data

Singer et al. (1987) studied nerve conduction velocity of the median motor, median sensor and

sural nerves in 16 workers exposed to various inorganic mercury compounds (e.g., mercuric oxides and

mercurial chlorides) for an average of 7.3 ± 7.1 years and compared to an unexposed control group using

t-tests. They found a slowing of nerve conduction velocity in motor, but not sensory, nerves that

correlated with increased blood and urine mercury levels and an increased number of neurologic

symptoms. The mean mercury levels in the exposed workers were 1.4 g/L and 10 g/L for blood and

urine, respectively. These urine levels are 2-fold less than those associated with peripheral neurotoxicity

in other studies (e.g., Levine et al. 1982). There was considerable variability in the data presented by

Singer et al. (1987), however, and the statistical analyses (t-test) were not as rigorous as those employed

by Levine et al. (1982) who used linearized regression analysis. Furthermore, the subsections in the

Levine et al. (1982) study were asymptomatic at higher urinary levels than those reported to be

associated with subjective neurological complaints in the workers studied by Singer et al. (1987). These

results, therefore, are not considered to be as reliable as those reported by Levine et al. (1982).
3-36

Kazantzis et al. (1962) performed renal biopsies in 2 (out of 4) workers with nephrotic syndrome

who had been occupationally exposed to mercuric oxide, mercuric acetate and probably mercury vapors. 

The authors felt that the nephrotic syndrome seen in 3 of the 4 workers may have been an idiosyncratic

reaction since many other workers in a factory survey had similarly high levels of urine mercury without

developing proteinuria. This conclusion was strengthened by work in Brown Norway rats indicating a

genetic (strain) susceptibility and that similar mercury-induced immune system responses have been seen

in affected humans and the susceptible Brown Norway rats (U.S. EPA 1987b).

3.2.1.2 Animal Data

Bernaudin et al. (1981) reported that mercurials administered by inhalation or ingestion to Brown

Norway rats resulted in the development of a systemic autoimmune disease. The mercuric chloride

ingestion portion of the study involved the forcible feeding of either 0 or 3 mg/kg/week of mercuric

chloride to male and female Brown Norway rats for up to 60 days. No abnormalities were reported using

standard histological techniques in either experimental or control rats. Immunofluorescence histology

revealed that 80% (4/5) of the mercuric-exposed rats were observed with a linear IgG deposition in the

glomeruli after 15 days of exposure. After 60 days of mercuric chloride exposure, 100% (5/5) of the rats

were seen with a mixed linear and granular pattern of IgG deposition in the glomeruli and granular IgG

deposition in the arteries. Weak proteinuria was observed in 60% (3/5) of the rats fed mercuric chloride

for 60 days. The control rats were observed to have no deposition of IgG in the glomeruli or arteries as

well as normal urine protein concentrations.

Andres (1984) administered mercuric chloride (3 mg per kg of body weight in 1 mL of water) by

gavage to five Brown Norway rats and two Lewis rats twice a week for 60 days. A sixth Brown Norway

rat was given only 1 mL of water by gavage twice a week for 60 days. All rats had free access to tap

water and pellet food. After 2–3 weeks of exposure, the Brown Norway mercuric chloride-treated rats

started to lose weight and hair. Two of the mercuric chloride-treated Brown Norway rats died 30–40

days after beginning the study. No rats were observed to develop detectable proteinuria during the 60-

day study. The kidneys appeared normal in all animals when evaluated using standard histological

techniques, but examination by immuno-fluorescence showed deposits of IgG present in the renal

glomeruli of only the mercuric-treated Brown Norway rats. The Brown Norway treated rats were also

observed with mercury-induced morphological lesions of the ileum and colon with abnormal deposits of

IgA in the basement membranes of the intestinal glands and of IgG in the basement membranes of the

lamina propria. All observations in the Lewis rats and the control Brown Norway rat appeared normal.

The only chronic oral study designed to evaluate the toxicity of mercury salts was reported by

Fitzhugh et al. (1950). In this study, rats of both sexes (20–24/group) were given 0.5, 2.5, 10, 40 or 160

ppm mercury as mercuric acetate in their food for up to 2 years. Assuming food consumption was equal

to 5% body weight per day, the daily intake would have been 0.025, 0.125, 0.50, 2.0 and 8.0 mg Hg/kg

for the five groups, respectively. At the highest dose level, a slight depression of body weight was

detected in male rats only. The statistical significance of this body weight depression was not stated. 

Kidney weights were significantly (p<0.05) increased at the 2-and 8-mg Hg/kg-day dose levels. 

Pathologic changes originating in the proximal convoluted tubules of the kidneys were also noted with

more severe effects in females than males. The primary weaknesses of this study were the lack of

reporting (which adverse effects were observed with which dosing groups) and that the most sensitive

strain, the Brown Norway rat, was not used for evaluating the mercury-induced adverse health effects.

NTP (1993) conducted subchronic and chronic gavage toxicity studies on Fischer 344 rats and

B6C3F1 mice to evaluate the effects of mercuric chloride, and the kidney appeared to be the major organ
3-37

of toxicity. These studies were also summarized by Dieter et al. (1992). In the 6-month study, Fischer

344 rats (10/sex/group) were administered 0, 0.312, 0.625, 1.25, 2.5, or 5 mg/kg-day of mercuric chloride

(0.23, 0.46, 0.92, 1.9, and 3.7 mg Hg/kg-day), 5 days/week, by gavage. Survival was not affected,

although body weight gains were decreased in males at high dose and in females at or above 0.46 mg

Hg/kg-day. Alkaline phosphatase and gamma-glutamyl transferase levels in the urine were significantly

elevated in the females exposed to 3.7 mg Hg/kg-day at four and six months of exposure. Absolute and

relative kidney weights were significantly increased in both sexes with exposure to at least 0.46 mg

Hg/kg-day. The kidney weight changes were slightly dose-related in the females. Histopathology

revealed corresponding changes in the kidneys. In males, the incidence of nephropathy was 80% in

controls and 100% for all treated groups; however, severity was minimal in the two low-dose groups and

minimal to mild in the 0.92-mg Hg/kg-day group and higher. In females, there was a significant

increased incidence of nephropathy only at the high-dose group (4/10 with minimal severity). 

Nephropathy was characterized by foci of tubular regeneration, thickened tubular basement membrane

and scattered dilated tubules containing hyaline casts. No treatment-related effects were observed in the

other organs; however, histopathology on the other organs was performed only on control and high-dose

rats.

B6C3F1 mice (10/sex/group) were administered gavage doses of 0, 1.25, 2.5, 5, 10, or 20 mg/kgday mercuric chloride (0, 0.92, 1.9, 3.7, 7.4, or 14.8 mg Hg/kg-day) 5 days/week for 6 months (NTP

1993). There was a decrease in body weight gain in males at the highest dose tested. Significant

increases occurred in absolute kidney weights of male mice dosed with 3.7 mg Hg/kg-day or more, and

relative kidney weights were increased in male mice at the 7.4 and 14.8 mg Hg/kg-day doses. The kidney

weight changes corresponded to an increased incidence of cytoplasmic vacuolation of renal tubule

epithelium in males exposed to at least 3.7 mg Hg/kg-day. The exposed female mice did not exhibit any

histopathologic changes in the kidneys.

In the 2-year NTP study, Fischer 344 rats (60 per sex per group) were administered 0, 2.5, and 5

mg/kg-day mercuric chloride (1.9 and 3.7 mg Hg/kg-day), 5 days a week, by gavage (Dieter et al. 1992;

NTP 1993). After two years, survival was significantly reduced in the treated male rats compared to the

controls. Mean body weights were significantly decreased in both treated males and females (9–10% and

14–15% decrease from control, respectively). At 15 months, relative kidney weights were significantly

elevated (not dose-related) in all treated groups (15–20% increase from control), and relative brain

weights were significantly elevated (slightly dose-related) in treated females (13–18%). The increased

kidney weights were accompanied by an increase in severity of nephropathy. After two years, there was

an increased incidence of nephropathy of moderate-to-marked severity and increased incidence of tubule

hyperplasia in the kidneys of exposed males compared to the controls. The control males exhibited

nephropathy, primarily of mild-to-moderate severity. Hyperparathyroidism, mineralization of the heart

and fibrous osteodystrophy were observed and considered secondary to the renal impairment. There

were no significant differences found in renal effects between exposed and control females. Other

nonneoplastic effects included an increased incidence of forestomach hyperplasia in the exposed males

and high dose females, increased incidence of nasal inflammation at the high-dose animals, slightly

increased incidence of acute hepatic necrosis in the high-dose males and increased incidence of

inflammation of the cecum in exposed males. Statistical analyses, however, were not performed on these

histopathologic changes.

NTP (1993) also administered to B6C3F1 mice (60/sex/group) daily oral gavage doses of 0, 5, or

10 mg/kg mercuric chloride (0, 3.7 and 7.4 mg Hg/kg-day), 5 days a week, by gavage for 2 years. 

Survival and body weights of mice were slightly lower in mercuric chloride treated mice compared to

controls. Absolute kidney weights were significantly increased in the treated males while relative kidney

weights were significantly increased in high-dose males and both low- and high-dose females. 
3-38

Histopathology revealed an increase in the incidence and severity of nephropathy in exposed males (mild

severity in low dose and moderate-to-marked severity in high dose) and females (minimal severity in low

dose and minimal-to-mild severity in high dose). Nephropathy was defined as foci of proximal

convoluted tubules with thickened basement membrane and basophilic cells with scant cytoplasm. Some

affected convoluted tubules contained hyaline casts. There was also an increase in nasal cavity

inflammation (primarily infiltration of granulocytes in nasal mucosa) in the exposed animals.

In a 4-week oral study (Jonker et al. 1993), Wistar rats (5–10/sex/group) were fed a diet

containing 15 and 120 ppm mercuric chloride (0.56 or 4.4 mg Hg/kg-day). A significant increase in

relative kidney weight was reported for the low-dose females and high-dose males. There was also an

increase in the incidence of high-dose males that had occasional basophilic tubules in the outer cortex of

the kidneys. In the range-finding study by Jonker et al. (1993), rats were administered 75, 150, or 300

ppm mercuric chloride (2.8, 5.6, 11.1 mg Hg/kg-day) in the diet for four weeks. A significant increase in

the relative kidney weights was observed in both sexes for all dose groups; the effect was dose related. 

Nephrosis and proteinaceous casts in the kidneys were reported in both sexes at the lowest dose. At 5.6

mg/kg-day, the body weight was significantly decreased in males and serum alkaline phosphatase levels

were elevated in females. At 11.1 mg/kg-day, increased serum aspartate aminotransferase (both sexes),

decreased urinary density (males), increased relative adrenal weight (males), increased serum sodium and

phosphate levels (females) and decrease in body weight (females) were reported.

A series of studies (Boscolo et al. 1989; Carmignani et al. 1989, 1992) reported renal and

cardiovascular changes in rats exposed to mercuric chloride in drinking water. These studies were

limited due to the small number of animals and dose levels tested. Boscolo et al. (1989) evaluated the

renal effects of mercuric chloride in two different rat strains. Male Sprague-Dawley rats (8/group) were

administered 0 or 0.05 mg/mL mercury (0 or 7 mg Hg/kg-day), and male Wistar rats (8/group) received

0, 0.05, or 0.2 mg/mL mercury (0, 7, or 28 mg Hg/kg-day) in drinking water for 350 days. Increases in

blood pressure and cardiac inotropism, without changes in heart rate, occurred in exposed rats of both

strains. Hydropic degeneration and desquamation of the proximal tubular cells were exhibited in kidneys

of Sprague-Dawley rats, with alterations and lysis of lysosomes in tubular cells and thickening of the

basal membrane in the glomeruli. Wistar rats displayed tubular degeneration and membranous

glomerulonephritis in 30% of the glomeruli at 7 mg/kg-day and all glomeruli at 28 mg/kg-day. 

Thickening of basal membrane and hypercellularity and alteration of the mesangial matrix in the

glomeruli and hydropic degeneration of tubules were seen in Wistar rats. Similar findings of renal

histopathology alterations and cardiovascular changes were reported by Carmignani et al. (1989) who

administered 0 or 0.05 mg/L of mercury (7 mg/kg-day) to male Sprague-Dawley rats (8/group) for 350

days.

In Carmignani et al. (1992), male Sprague-Dawley rats (8/group) received 0 or 0.2 mg/mL of

mercury (28 mg Hg/kg-day) as mercuric chloride in drinking water for a shorter duration (180 days). 

Similar renal changes were observed, as well as IgM deposition in the glomeruli (as shown by

immunofluorescence). In addition, the treated group displayed significantly decreased urinary kallikrein

and creatinine, decreased plasma renin and increased plasma angiotensin-converting enzyme. The

cardiovascular effects were slightly different from Boscolo et al. (1989) and Carmignani et al. (1989);

there was an increase in blood pressure but a decrease in cardiac inotropism in the exposed rats. The

increase in blood pressure was suggested to be due to a vasoconstrictor effect, likely related to a greater

release of noradrenaline from adrenergic neurons and to baroreflex hyposensitivity. The decrease in

contractility was attributed to a direct toxic effect of the mercury on the cardiac muscle because of the

high levels of mercury detected in the heart. The differences in the results of cardiovascular changes for

the studies were not explained.
3-39

To evaluate the effect of mercuric chloride on the development of autoimmunity, female SJL/N

mice (7/group) received 0.625, 1.25, 2.5, or 5 ppm mercuric chloride (0.07, 0.14, 0.28, or 0.56 mg Hg/kgday) in drinking water ad libitum for 10 weeks (Hultman and Enestrom 1992). An increase in circulating

antinucleolar antibodies was observed at 0.28 mg Hg/kg-day. The high-dose group had elevated granular

IgG deposits in the renal mesangium and in vessel walls of glomerular capillaries, arteries and arterioles

of the spleen and in intramyocardial arteries. Slight glomerular cell hyperplasia and discrete widening of

the centrolobular zone were also exhibited in the 0.56-mg/kg-day group.

Agrawal and Chansouria (1989) administered 0, 2.6, 5.2, and 10.4 mg Hg/kg-day as mercuric

chloride in drinking water to male Charles Foster rats for 60, 120, or 180 days (5/group). The relative

adrenal gland weight was significantly increased for the dose groups at all durations compared to

controls. Significant increases in adrenal and plasma corticosterone levels occurred in all dose groups at

60 and 120 days; however, changes were not seen after 180 days. The authors suggested that mercuric

chloride may have acted as a chemical stressor in a dose- and duration-dependent manner. The study was

limited because histopathology was not performed on the kidneys, and the adrenal gland was the only

tissue evaluated.

Both male and female Brown Norway rats 7–9 weeks of age were divided into groups of 6–20

animals each (Druet et al. 1978). The numbers of each sex were not stated. The animals were injected

subcutaneously with mercuric chloride 3 times weekly, for 8 weeks, with doses of 0, 0.07, 0.2, 0.4, 0.7,

and 1.5 mg Hg/kg. An additional group was injected with a 0.04 mg/kg for 12 weeks. Antibody

formation was measured by the use of kidney cryostat sections stained with a fluoresceinated sheep antirat IgG antiserum; urinary protein was assessed by the biuret method. Tubular lesions were seen at the

higher dose levels. Proteinuria was seen at doses of 0.07 mg/kg and above, but not at 0.04 mg/kg. 

Proteinuria was considered a highly deleterious effect in that affected animals developed

hypoalbuminemia and many died. Fixation of IgG antiserum was detected in all groups except controls.

3.2.2 Cancer Data

3.2.2.1 Human data

No data are available on the carcinogenic effects of inorganic mercury in humans. 

3.2.2.2 Animal data

The results from a dietary study in rats and mice show equivocal evidence for carcinogenic

activity in male mice and female rats and some evidence for carcinogenic activity in male rats. Two

other dietary studies show negative evidence for carcinogenicity, but these studies are limited by

inadequacies in the data and experimental design.

Mercuric chloride was administered by gavage in water at doses of 0, 2.5, or 5 mg/kg-day (0, 1.9

and 3.7 mg Hg/kg-day) to Fischer 344 rats (60/sex/group), 5 days a week, for over 104 weeks (NTP

1993). An interim sacrifice (10/sex/dose) was conducted after 15 months of exposure. Complete

histopathological examinations were performed on all animals found dead, killed in extremis, or killed by

design. Survival after 24 months was statistically significantly (p<0.01) lower in low- and high-dose

males; survival was 43%, 17% and 8% in control, low-, and high-dose males, respectively, and 58%,

47%, and 50% in control, low-, and high-dose females, respectively. During the second year of the

study, body weight gains of low- and high-dose males were 91% and 85% of controls, respectively, and

body weight gains of low- and high-dose females were 90% and 86% of controls, respectively. At study

termination, nephropathy was evident in almost all male and female rats including controls, but the
3-40

severity was much greater in treated males; the incidence of "marked" nephropathy was 6/50, 29/50, and

29/50 in control, low- and high-dose males, respectively. 

Squamous cell papillomas of the forestomach showed a statistically significant (p<0.001)

positive trend with dose by life table adjusted analysis; the incidences were 0/50, 3/50 and 12/50 in

control, low-, and high-dose males, respectively. The incidence in female rats was 0/50, 0/49 and 2/50 in

control, low- and high-dose groups, respectively. These neoplasms are rare neoplasms in rats and

occurred in only 1 out of 264 historical controls. The incidence of papillary hyperplasia of the stratified

squamous epithelium lining of the forestomach was statistically significantly (p0.01) elevated in all

dosed males (3/49, 16/50 and 35/50 in control, low- and high-dose males, respectively) and in high-dose

females (5/50, 5/49 and 20/50 in control, low- and high-dose females, respectively). The incidence of

thyroid follicular cell carcinomas was marginally significantly (p=0.044 by logistic regression analysis;

tumors not considered to be fatal) increased in high-dose males (1/50, 2/50 and 6/50 in control, low- and

high-dose groups, respectively). The data, adjusted for survival, also showed a significant (p=0.017)

positive trend in males. The combined incidence of thyroid follicular cell neoplasms (adenoma and/or

carcinoma), however, was not significantly increased (2/50, 6/50 and 6/50 in control, low- and high-dose

males, respectively). In female rats a significant decrease in the incidence of mammary gland

fibroadenomas was observed (15/50, 5/48 and 2/50 in control, low- and high-dose females, respectively). 

Table 3-31 gives the incidences of lesions which were increased in treated animals.

The high mortality in both groups of treated males indicates that the maximally tolerated dose

(MTD) was exceeded in these groups and limits the interpretation of the study. NTP (1993) considered

the forestomach tumors to be of limited relevance to humans because the tumors did not appear to

progress to malignancy. NTP (1993) also questioned the relevance of the thyroid carcinomas because

these neoplasms are usually seen in conjunction with increased incidences of hyperplasia and adenomas,

but increases in hyperplasia (2/50, 4/50 and 2/50 in control, low- and high-dose males, respectively) or

adenomas (1/50, 4/50 and 0/50 in control, low- and high-dose males, respectively) were not observed.

In the same study, mercuric chloride was administered by gavage in water at doses of 0, 5, or

10 mg/kg-day (0, 3.7 and 7.4 mg Hg/kg-day), 5 days a week, for 104 weeks to B6C3F1 mice

(60/sex/group) (NTP 1993). An interim sacrifice (10/sex/dose) was conducted after 15 months of

exposure. Terminal survival of male mice was not affected by the administration of mercuric chloride;

survival of high-dose females was slightly lower (p=0.051) than controls (41/60, 35/60 and 31/60 in

control, low- and high-dose females, respectively). Body weight gain was not affected. Female mice

exhibited a significant increase in the incidence of nephropathy (21/49, 43/50 and 42/50 in control, lowand high-dose females, respectively). Nephropathy was observed in 80–90% of the males in all groups. 

The severity of nephropathy increased with increasing dose (1.08, 1.74 and 2.51 in control, low- and

high-dose males, respectively; 0.47, 1.02 and 1.24 in control, low- and high-dose females, respectively). 

The incidence of renal tubule hyperplasia was 1/50, 0/50 and 2/49 in control, low- and high-dose males.
3-41

Table 3-31

Incidence of Selected Lesions in Rats in the NTP (1993) 2-Year Gavage Study a

Tumor Site and Type Dose Group (mg Hg/kg-day)

Males Females

0 1.9 3.7 0 1.9 3.7

Forestomach

 Papillary hyperplasia 3/49 16/50 35/50 5/50 5/49 20/50 b b b

 Squamous cell papilloma 0/50 3/50 12/50 0/50 0/49 2/50 c

Thyroid Follicular Celld

 Adenoma 1/50 4/50 0/50 -- -- --

 Carcinoma 1/50 2/50 6/50 -- -- -- e

 Adenoma or carcinoma 2/50 6/50 6/50 -- -- --

 Overall rate a

 p  0.01 b

 p 0.001; trend test also p<0.001 c

 Data on thyroid follicular cell lesions were reported for males only. d

 p = 0.044, logistic regression e

As shown in Table 3-32, the combined incidence of renal tubule adenomas and adenocarcinomas

was 0/50, 0/50 and 3/49 in control, low- and high-dose males, respectively. Although no tumors were

seen in the low-dose group, a statistically significant (p=0.032) positive trend for increased incidence

with increased dose was observed. These observations were considered important because renal tubule

hyperplasia and tumors in mice are rare. The two-year historical incidence of renal tubule adenomas or

adenocarcinomas in male mice dosed by gavage with water was 0/205, and only four of the nearly 400

completed NTP studies have shown increased renal tubule neoplasms in mice. NTP did not report a

statistical comparison of the study data to historical control data. Analysis of the reported data with

Fisher's Exact test, however, showed that the incidence of renal tubule adenomas or adenomas and

carcinomas (combined) in the high-dose males was significantly elevated when compared to historical

controls (Rice and Knauf 1994).
3-42

Table 3-32

Incidence of Renal Tubule Tumors in Male Mice in the a

NTP (1993) 2-Year Gavage Study

Dose Group (mg/kg-day)

0 5 10

Adenoma 0/50 0/50 2/49

Adenomacarcinoma 0/50 0/50 1/49

Adenoma or adenomacarcinoma 0/50 0/50 3/49b

 Overall rate a

 p = 0.107; trend test p = 0.032 b

 A 2-year feeding study in rats (20 or 24/sex/group; strain not specified) was conducted in which

mercuric acetate was administered in the diet at doses of 0, 0.5, 2.5, 10, 40, and 160 ppm (0, 0.02, 0.1,

0.4, 1.7, and 6.9 mg Hg/kg-day) (Fitzhugh et al. 1950). Survival was not adversely affected in the study. 

Increases in kidney weight and renal tubular lesions were observed at the two highest doses. No

statement was made in the study regarding carcinogenicity. This study was not intended as a

carcinogenicity assay, and the number of animals/dose was rather small. Histopathological analyses

were conducted on only 50% of the animals (complete histopathological analyses were conducted on

only 31% of the animals examined), and no quantitation of results or statistical analysis was performed.

No increase in tumor incidence was observed in a carcinogenicity study using white Swiss mice

(Schroeder and Mitchener 1975). Groups of mice (54/sex/group) were exposed until death to mercuric

chloride in drinking water at 5 ppm mercury (0.95 mg Hg/kg-day). No effects on survival or body

weights were observed. After dying, mice were weighed, dissected, gross tumors were detected, and

some sections were made of the heart, lung, liver, kidney and spleen for microscopic examination. 

Mercuric chloride was nontoxic in the study. No statistically significant differences were observed in

tumor incidences for treated animals and controls. This study is limited because complete histological

examinations were not performed, only a single dose was tested, and the MTD was not achieved.

The increasing trend for renal tubular cell tumors in mice observed in the NTP (1993) study is

supported by similar findings in mice after chronic dietary exposure to methylmercury (Hirano et al.

1986; Mitsumori et al. 1981, 1990). In these studies, dietary exposure to methylmercuric chloride

resulted in increases in renal tubular tumors at doses where substantial nephrotoxicity was observed.
3-43

Table 3-33

Carcinogenic Effects of Inorganic Mercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations//BML Reference

Rat/strain 2 yr 0, 0.02, 0.1, No carcinogenicity reported Fitzhugh et al. 1950

NS/20–24 M, ad lib 0.4, 1.7, 6.9 Limitations: small number of animals/dose; complete

20-24 F histopathological examinations conducted on only 31% of

a

animals; no statistical analyses

Rat/F344/ 2 yr 0, 1.9, 3.7 Thyroid follicular cell carcinomas in males at 3.7. NTP 1993

60 M, 60 F 5 d/wk (HgCl ) Limitations: MTD exceeded (high mortality in treated

1 x/d males); limited relevance of lesions to humans

(gavage)

2

Mouse/Swiss/54 Lifetime 0.95 No carcinogenicity reported Schroeder and

M, 54 F ad lib (HgCl ) Limitations: Complete histopathological examinations not Mitchener 1975 2

performed; only single dose level tested; MTD not

achieved

Mouse/ 2 yr 0, 3.7, 7.4 Renal tubule tumors (adenoma or adenomacarcinoma) in NTP 1993

B6C3F1/ 5 d/wk (HgCl ) 3/49 males at 7.4 (positive trend test, p=0.0032)

60 M, 60 F 1 x/d Limitations: Severe nephropathy also observed in high-

(gavage) dose males.

2

 Phenylmercuric acetate and mercuric acetate a

3.2.3 Other Data

3.2.3.1 Death

The estimated lethal dose of inorganic mercury for a 70 kg adult is 10–42 mg Hg/kg (Gleason et

al. 1957). Most deaths attributed to inorganic mercury occur soon after a person ingests a single large

amount of mercury. Causes of death include cardiovascular failure, gastrointestinal damage and acute

renal failure (Troen et al. 1951).

Table 3-34

Lethality of Inorganic Mercury in Humans: Case Study

Species/

No. per Exposure Dose

Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/25 M, Once 21-37 (est.) Case studies of mercuric chloride poisonings in victims age Troen et al. 1951

29 F (HgC1 ) 2-60 yr; 9 resulted in death (all adults). 2

BML not reported
3-44

The estimated LD for rats following oral exposure to mercuric chloride is 25.9 mg Hg/kg; 50

however, LD levels as high as 77.7 mg Hg/kg have been observed in rats (Kostial et al. 1978). Male 50

rats appear to be more sensitive to the effects of mercuric chloride. This was demonstrated in a chronicduration oral study with rats, in which 40/50 males and 21/49 females died at the low dose, 45/50 males

and 20/50 females died at the high dose, compared to 24/50 males and 15/50 females in the controls

(Dieter et al. 1992; NTP 1993). The increase in deaths in the male rats was statistically significant and

were considered to be due to renal lesions. Mortality incidence was not significantly increased in

exposed female groups.

Table 3-35

Lethality of Inorganic Mercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/Albino Once NS LD = 25.8 mg/kg for 2-week old pups; older rats had Kostial et al. 1978

(NS)/6 NS (gavage) (6 levels) higher LD values.

(HgCl ) Limitation: Incomplete data reporting (i.e., doses not 2

50

50

reported, toxic effects not specified)

BML not reported

Rat/F344/5 M, 14 d 0, 0.93, 1.9, 2/5 males died at 14.8; no other animals died. Dieter et al. 1992

5 F 5 d/wk 3.7, 7.4, 14.8 BML: 45.4 μg/g in kidney of males, 43.3 μg/g in kidney

1 x/d (HgCl ) of females

(gavage) 2

Rat/F344/ 2 yr 0, 1.9, 3.7 40/50 males died at 1.9 mg/kg, vs. 24/50 control males; Dieter et al. 1992

50 M, 50 F 5 d/wk (HgCl ) survival of dosed females was not significantly different

1 x/d from controls.

(gavage) BML not reported

2

Mouse/ 14 d 0, 3.7, 7.4, 9/10 died (LOAEL = 59). NTP 1993

B6C3F /5 M, 5 5 d/wk 14.8, 29, 59 Limitations: small number of animals 1

F 1 x/d (HgCl ) BML not reported

(gavage) 2

3.2.3.2 Neurological

Limited studies are available concerning neurological toxicity following oral exposure to

inorganic mercury. These studies are summarized below.
3-45

Table 3-36

Neurotoxicity of Inorganic Mercury in Humans: Case Studies

Species/ Exposure Dose

No. per Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/2 F 6–25 yr 0.73 Dementia, irritability, decreased cerebellar neurons, low Davis et al. 1974

(Hg Cl ) brain weight 2 2

Limitation: Case study

BML: 3.4-4.7 μg/g in frontal cortex

Human/2 M (1 3 mo NS Drooling, dysphagia, irregular arm movements, impaired Kang-Yum and

child), 1 adult F (Hg Cl ) gait, convulsions following ingestion of patent medicines Oransky 1992 2 2

(HgS) containing mercuric sulfide and mercurous chloride

Limitation: Case studies; concomitant exposure to other

metals; limited exposure data

BML: 39-2800 μg/L in 24 hr urine

There are several animal studies in which inorganic mercury-induced neurotoxicity has been

reported.

Table 3-37

Neurotoxicity of Inorganic Mercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/ 11 wk 0, 0.74 Weakening of hind legs, crossing reflex of limbs, ataxia; Chang and

Holtzman/ (HgCl ) degenerative changes in neurons of dorsal root ganglia and Hartmann 1972

8 M exposed, Purkinje and granule cells of cerebellum

8 M control Limitation: One dose level tested

2

BML not reported

Rat/Sprague 3 mo ad lib in 0, 2.2 Inactivity and abnormal gait Goldman and

Dawley/12 F feed (HgCl ) Limitation: One dose level tested Blackburn 1979

exposed, 10 F BML not reported

controls

2

Mouse/C57BL 17 mo 0.74 for 110 d, No clinical signs of neurotoxicity; no effect on optic or Ganser and

6J/NS ad lib in then 7.4–14.8 peripheral nerve structure Kirschner 1985

drinking for 400 d; 2.2 Limitation: Lack of statistical analyses due to insufficient

water for 17 mo number of animals tested; uncertainty of dosage due to

(HgCl ) large variation in water consumption 2

BML not reported

3.2.3.3 Renal

The kidney appears to be the critical target organ for the effects of acute ingestion of inorganic

mercury. Case studies of poisonings by mercuric chloride report acute renal failure, including
3-46

proteinuria, oliguria and hematuria, in people ingesting estimated doses of 3.5–37 mg Hg/kg (Afonso and

deAlvarez 1960; Pesce et al. 1977; Troen et al. 1951). These effects are attributed to tubular and

glomerular pathology.

Table 3-38

Renal Toxicity of Inorganic Mercury in Humans: Case Studies

Species/ Exposure Dose

No. per Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/25 M, Once 3.5-37 (est.) Case studies of mercuric chloride poisonings in victims Troen et al. 1951

29 F (HgCl ) age 2-60 yr; 18 cases resulted in renal effects 2

(albuminuria, anuria)

BML not reported

Human/1 F Once 30 Oliguria; proteinuria; hematuria following ingestion of Afonso and

(adult) (tablet) (HgCl ) mercuric chloride deAlvarez 1960 2

Limitation: Case study

BML not reported

Human/1 M Once 21.4 Proteinuria indicating glomerular and tubular damage Pesce et al. 1977

(HgCl ) Limitation: Case study 2

BML: Avg 370 μg/L in blood

There are numerous animal studies reporting kidney damage in rats and mice ingesting inorganic

mercury. Acute exposures result in increased kidney weight with at least 0.46 mg Hg/kg-day and tubular

necrosis at higher doses; males appear to have greater sensitivity for the histological changes than

females (Fowler 1972; NTP 1993). Similarly, longer-term studies have found histopathologic effects

affecting the tubules and glomeruli, including thickening of basement membranes and degeneration of

tubular cells (Carmignani et al. 1989; Jonker et al. 1993; NTP 1993). A study monitoring kidney

function reported ketonuria and proteinaceous casts (Jonker et al. 1993).

Table 3-39

Renal Toxicity of Inorganic Mercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/Sprague 350 d 0, 7 Hydropic degeneration of tubular cells Carmignani et al.

Dawley/8 M ad lib in (HgCl ) Limitation: Only one dose tested 1989

drinking BML: 140 μg/g in kidney

water

2
Table 3-39 (continued)

Renal Toxicity of Inorganic Mercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

3-47

Rat/Sprague- 180 d 0, 28 Hydropic degeneration of tubular cells, IgM deposition Carmignani et al.

Dawley/8 M ad lib in (HgCl ) in glomeruli, decreased urinary kallikrein and 1992

drinking creatinine, decreased plasma renin, increased plasma

water angiotensin-converting enzyme

2

Limitation: Only one dose tested

BML: 0.94 μg/g in blood

Rat/Wistar/5 M, 5 4 wk 0, 0.56, 4.4 Ketonuria in males at 4.4 mg/k-day; increased kidney Jonker et al. 1993

F exposed/10 M, ad lib in feed (HgCl ) weight in males and females (LOAEL = 0.56 in

10 F controls females, 4.4 in males) 2

BML not reported

Rat/Wistar/ 4 wk 2.8, 5.6, 11.1 Ketonuria in males at all levels; increased relative Jonker et al. 1993

5 M, 5 F ad lib in feed (HgCl ) kidney weight, increased nephrosis and proteinaceous

exposed/10 M, 10 casts in males and females (LOAEL = 2.8)

F controls BML not reported

2

Rat/F344/ 14 d 0, 0.93, 1.9, Increased absolute and relative kidney weight in males NTP 1993

5 M, 5 F 5 d/wk 3.7, 7.4, 14.8 and females (LOAEL = 1.9); acute renal tubule necrosis

1 x/d (HgCl ) at 3.7 mg/kg-day in both sexes

(gavage) BML: 43-46 μg/g in kidney at 14.8 mg/kg 2

Rat/F344/ 6 mo 0, 0.23, 0.46, Increased absolute and relative kidney weight in males NTP 1993

10 M, 10 F 5 d/wk 0.93, 1.8, 3.7 and females (LOAEL = 0.46); increased severity of

1 x/d (HgCl ) nephropathy in males (LOAEL = 0.93)

(gavage) BML: 86.2-89.6 μg/g in kidney at 0.93 mg/kg 2

Rat/F344/ 2 yr 0, 1.9, 3.7 Increased severity of nephropathy in males (thickening NTP 1993

60 M, 60 F 5 d/wk (HgCl ) of glomerular and tubular basement membranes;

1 x/d degeneration and atrophy of tubule epithelium)

(gavage) (LOAEL = 1.9)

2

BML not reported

Mouse/NMRI/20 Once 0, 5, 10, 20, 40 Decreased selenium-dependent glutathione peroxidase Nielsen et al. 1991

(sex NS) (gavage) (HgCl ) activity in kidney; minor renal tubular damage (LOAEL

exposed/10 = 10)

controls BML: 260 μg/L in blood at 10 mg/kg

2

Mouse/NMRI/24 Once 0, 20 Necrosis of proximal tubules Nielsen et al. 1991

(sex NS) (gavage) (HgCl ) BML not reported 2

Mouse/B6C3F / 14 d 0, 3.7, 7.4, Increased absolute and relative kidney weight (LOAEL NTP 1993 1

5 M, 5 F 5 d/wk 14.8, 29, 59 = 3.7); acute renal tubular necrosis at 29 mg/kg-day in

1 x/d (HgCl ) males and at 59 mg/kg-day in males and females

(gavage) BML: 116-171 μg/g in kidney at 29 mg/kg-day 2

Mouse/B6C3F / 6 mo 0, 0.93, 1.9, Increased absolute and relative kidney weight of males NTP 1993 1

10 M, 10 F 5 d/wk 3.7, 7.4, 14.8 (LOAEL = 3.7); Cytoplasmic vacuolation of tubule

1 x/d (HgCl ) epithelium in males (LOAEL = 3.7)

(gavage) BML: 36.1-40.6 μg/g in kidney at 3.7 mg/kg-day 2
Table 3-39 (continued)

Renal Toxicity of Inorganic Mercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

3-48

Mouse/B6C3F / 2 yr 0, 3.7, 7.4 Increased severity of nephropathy (foci of proximal NTP 1993 1

60 M, 60 F 5 d/wk (HgCl ) tubules with thickened basement membrane; basophilic

1 x/d cells with scant cytoplasm (LOAEL = 3.7)

(gavage) BML not reported

2

Bernaudin et al. (1981) exposed male and female Brown Norway rats (number not specified) to

mercuric chloride via aerosols (4 hours/week) and intratracheal instillation for 2 months. The aerosol

exposures resulted in a retention of 0.05–0.06 mg HgCl /kg/hour (based on radiolabeled mercury); the 2

parameters of the aerosol were not well characterized (e.g., no mass median aerodynamic diameter

(MMAD) or geometric standard deviation provided and the particle generation system was not

adequately described). The autoimmune response was typified by a linear pattern of IgG conjugate

fixation in kidney glomeruli and a granular pattern of fixation in kidney glomeruli and arteries, lung and

spleen; evidence of autoimmune disease was noted at all but the lowest intratracheal exposure level (60

μg HgCl /kg/week). In two of three rats exposed to aerosols and examined when sacrificed, weak 2

proteinuria (1, 28 and 47 mg/day) was detected. No significant proteinuria was observed in the animals

administered mercuric chloride by intratracheal instillations.

Table 3-40

Renal Toxicity of Inorganic Mercury in Animals: Inhalation Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Rat/Brown 2 mo, 1x/wk 0, 6, 11, 47, Autoimmune effect in spleen at 6 mg/kg-day and in Bernaudin et al.

Norway/ 3-8 both (intra- 79 mg/kg-day spleen, lung and kidney at higher doses 1981

sexes tracheal) (HgCl ) BML not reported 2

Rat/Brown 2 mo, 1 Weak proteinuria; autoimmune effect in kidney, lung Bernaudin et al.

Norway/5 both 4 d/wk, (HgCl /m ) and spleen 1981

sexes 1 hr/d (estimate of BML not reported

(aerosol) minimum air

2

3

concentration)

3.2.3.4 Cardiovascular

No studies were located regarding the cardiovascular toxicity of inorganic mercury in humans

following oral exposure.
3-49

Limited information was located regarding the cardiovascular toxicity of inorganic mercury

following oral exposure in animals. Signs of cardiovascular toxicity in rats include increased blood

pressure and varying changes in the contractility of the heart (Carmignani et al. 1989, 1992). These signs

manifested after oral exposure to mercuric chloride in drinking water for 180 or 350 days. No other

animal studies were located.

Table 3-41

Cardiovascular Toxicity of Inorganic Mercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/Sprague 350 d 0, 7 Increased blood pressure; positive inotropic response Carmignani et al.

Dawley/8 M ad lib in (HgCl ) (p<0.05) 1989

drinking Limitation: Only one dose tested; small number of

water animals

2

BML: 0.9 μg/g in heart

Rat/Wistar/8 M 180 d 0, 28 Increased blood pressure (p<0.05); negative inotropic Carmignani et al.

ad lib in (HgCl ) response (not significant) 1992

drinking Limitation: Only one dose tested; small number of

water animals

2

BML: 940 μg/L in blood, 4.1 μg/g in heart

3.2.3.5 Gastrointestinal

Irritation of the gastrointestinal mucosa is a common outcome of mercury toxicity following

ingestion of mercuric chloride (Murphy et al. 1979). Ingestion of inorganic mercury may also cause

vomiting, nausea, severe abdominal pain and diarrhea (Afonso and deAlvarez 1960; Murphy et al. 1979). 

No studies were located regarding the gastrointestinal toxicity of inorganic mercury after ingestion in

humans for intermediate or chronic durations.

Table 3-42

Gastrointestinal Toxicity of Inorganic Mercury in Humans: Case Studies

Species/ Exposure Dose

No. per Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/25 M, Once 3.5-37 (est.) Case studies of mercuric chloride poisonings in victims Troen et al. 1951

29 F (HgCl ) age 2-60 yr; effects ranged from nausea to severe corrosive 2

gastritis

Limitation: exposure data limited

BML not reported

Human/1 F Once 30 Nausea; vomiting; abdominal cramps; diarrhea after Afonso and

(adult) (tablets) (HgCl ) ingestion of mercuric chloride deAlvarez 1960 2

Limitation: Case study

BML not reported
3-50

Similar signs of gastrointestinal irritation appear in mice after intermediate duration oral

exposure to mercuric chloride (NTP 1993). Histopathologic analyses reveal inflammation and necrosis

of the stomach tissue. Further damage occurs to the gastrointestinal tract with continued dosing (NTP

1993). The incidence of hyperplasia of the forestomach epithelium increases in high-dose rats fed

mercuric chloride for two years (NTP 1993).

Table 3-43

Gastrointestinal Toxicity of Inorganic Mercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/F344/60 2 yr 0, 1.9, 3.7 Forestomach epithelial hyperplasia (LOAEL = 1.9 in NTP 1993

M, 60 F 5 d/wk (HgCl ) males, 3.7 in females).

1 x/d BML not reported

(gavage)

2

Mouse/ 14 d 0, 3.7, 7.4, Stomach inflammation and necrosis (LOAEL = 59). NTP 1993

B6C3F /5 M, 5 5 d/wk 14.8, 29, 59 BML: 116-171 μg/g in kidneys at 29 mg/kg-day 1

F 1 x/d (HgCl ) Limitation: small number of animals

(gavage) 2

3.2.3.6 Hepatic

Limited information is available regarding the hepatic toxicity of inorganic mercury in humans

after oral exposure. Murphy et al. (1979) reported on a man who died after ingesting an unspecified

amount of mercuric chloride. The man was jaundiced, and his liver enzymes were elevated prior to

death. Histopathological analyses revealed a softened and enlarged liver.

Liver enzymes were increased in rats and mice that ingested mercuric chloride for 4 or 6 weeks

(Dieter et al. 1983; Jonker et al. 1993; Rana and Boora 1992). In addition, liver weights were

significantly increased in mice after exposure to mercuric chloride in drinking water (Dieter et al. 1983). 

No microscopic changes were seen, however, during the histopathological analyses. No other animal

data were available regarding oral exposure to inorganic mercury.
3-51

Table 3-44

Hepatic Toxicity of Inorganic Mercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/Charles 30 d NS Increased lipid peroxidation (p<0.02) Rana and Boora

Foster/5 M 1 x/d (HgCl ) Limitation: Only one dose (NS) tested 1992

feed BML not reported 2

Rat/Wistar/ 4 wk 2.8, 5.6, 11.1 Increased serum alkaline phosphatase in males and females Jonker et al. 1993

5/sex ad lib in feed (HgCl ) (LOAEL = 5.6 in females, 11.1 in males)

exposed/10/sex Limitation: small number of animals

controls BML not reported

2

Mouse/ 1-7 wk 0, 0.6, 2.9, 14.3 Increased plasma cholinesterase (LOAEL = 2.9) Dieter et al. 1983

B6C3F /10 M ad lib in (HgCl ) Limitation: small number of animals tested 1

drinking BML: 0.6 μg/L in blood at 7 wk, at 2.9 mg/kg/d

water

2

3.2.3.7 Immunological

In addition to the inorganic mercury-induced autoimmune glomerulonephritis discussed earlier

(see discussion of renal effects in Section 3.2.1), several studies identified other immunotoxicity

endpoints in animals after oral exposure to inorganic mercury.

3.2.3.8 Developmental

No studies were located regarding the developmental toxicity of inorganic mercury in humans

after inhalation exposure. 

The only information located regarding developmental toxicity in animals from inhalation

exposure to mercuric mercury comes from a study in which mice were exposed to aerosols containing

mercuric chloride during gestation (Selypes et al. 1984). Increases were observed in the incidence of

delayed ossification and dead or resorbed fetuses; the statistical significance of these effects was not

reported. In addition, at the highest concentration, a significant increase in weight retardation was also

observed. Interpretation of this study is limited, however, because the aerosols were not well

characterized, and it is not known to what extent the droplets were respirable or were cleared from the

upper respiratory tract and swallowed.
3-52

Table 3-45

Immunotoxicity of Inorganic Mercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/Brown- 2 mo 2.2 IgG deposits in glomerular capillary wall of kidney and Bernaudin et al.

Norway/6 both 1 x/wk (HgCl ) renal arteries, suggestive of autoimmune disease; similar 1981

sexes (gavage) deposits also observed in lungs and spleen; no deposits

exposed/22 observed in controls

controls both Limitation: Only one dose tested; small number of

sexes animals tested

2

BML not reported 

Mouse/ 7 wk 0, 0.6, 2.9, 14.3 Suppression of lymphoproliferative response to T-cell, Dieter et al. 1983

B6C3F /10 M ad lib in (HgCl ) concavalin A and phytohemagglutinin (LOAEL = 2.9 1

drinking mg/kg-day; p<0.05)

water BML: 600 μg/L in blood at 2.9 mg/kg-day

2

Mouse/SJL or 2 wk 0, 0.7 Increased lymphoproliferative response to concanavalin A Hultman and

DBA/5 F ad lib in (HgCl ) and E. coli lipopolysaccharide (p<0.02) Johansson 1991

drinking Limitations: only one dose tested; small number of

water animals tested

2

BML not reported

Mouse/SJL/7 F 10 wk 0, 0.07, 0.14, Increased antinucleolar antibodies in IgG class (LOAEL = Hultman and

ad lib in 0.28, 0.56 0.28, p<0.05) Enestrom 1992

drinking (HgCl ) Limitation: small number of animals tested

water BML: 5.2 μg/g in kidney 2

Mouse/ 14 d 0, 3.7, 7.4, Decreased thymus weight (LOAEL = 14.8) NTP 1993

B6C3F /5 M, 5 5 d/wk 14.8, 29, 59 Limitation: small number of animals tested 1

F 1 x/d (HgCl ) BML: 116-171 μg/g in kidney at 29 mg/kg-day

(gavage) 2

Table 3-46

Developmental Toxicity of Inorganic Mercury in Animals: Inhalation Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Mice/CFLP/ 4 d 0, 0.17, 1.6 Increased dead or resorbed fetuses; delayed ossification Selypes et al. 1984

No. F NS 4 hr/d (HgCl ) (LOAEL = 0.17)

Gd 9-12 Limitations: Data were reported as number of embryos only, 2

not as number of affected litters; no statistical analysis; aerosol

exposure was not well characterized; maternal toxicity was not

evaluated

BML not reported
3-53

Developmental effects have been reported in animals following oral exposure to inorganic

mercury. These efforts include an increased incidence of abnormal fetuses in hamsters (Gale 1974),

growth retardation in rats (Rizzo and Furst 1972) and decreased body weights in several rat studies. 

Gale (1974) administered 0, 4, 8, 25, 35, 50, 75, or 100 mg mercuric acetate/kg (0, 2.5, 5, 16, 22,

32, 47, or 63 mg Hg/kg) to pregnant golden hamsters (10/exposed group; 3/control group) by gavage in

distilled water on the 8th day of gestation. The pregnant animals were sacrificed on gestation day 12 or

14, and the uterine contents were examined. A statistically significant increase in the incidence of

abnormal fetuses (combined incidence of small, retarded, edematous, and/or malformed fetuses) was

observed at 16 mg Hg/kg. Statistically significant increases in the percentage of resorbed fetuses was

observed at 22 mg Hg/kg and in the percentages of small, retarded and edematous fetuses observed at 32

mg Hg/kg. No treatment-related effects were observed on the fetuses at 5 mg Hg/kg. Toxic effects

observed in maternal animals included weight loss, diarrhea, slight tremor, somnolence, tubular necrosis

in the kidneys and cytoplasmic vacuolization of hepatocytes.

Rizzo and Furst (1972) administered 7 mg Hg/kg as mercuric oxide to pregnant Long-Evans

rats (5/group) by gavage in peanut oil on gestation day 5, 12, or 19 in a pilot study. On gestation day 20

or 21, the rats were sacrificed, and the uterine contents were examined. Rats administered mercury on

gestation day 5 had a higher percentage of fetuses with growth retardation and inhibition of eye

formation (statistical significance not reported). Similar increases in these effects were not observed

after administration on gestation day 12 or 19. No toxicity in maternal animals was reported.

McAnulty et al. (1982) administered 8, 12, 16, or 24 mg mercuric chloride/kg-day (6, 9, 12, or 18

mg Hg/kg-day) by gavage to pregnant rats (strain and number not specified) on gestation days 6–15 as

reported in an abstract. The abstract did not report whether controls were used. Fetal and placental

weights were decreased at 9 mg Hg/kg-day and above. At 12 and 18 mg Hg/kg-day, increased

postimplantation losses were reported. These effects were attributed to maternal toxicity and decreased

food intake. At 18 mg Hg/kg-day, increases in delayed ossification and malformations were reported. 

Statistical analyses were not reported.

Pritchard et al. (1982a) administered 4, 8, or 16 mg mercuric chloride/kg-day (3, 6, or 12 mg

Hg/kg-day) by the oral route to pregnant rats (number and strain not specified) from gestation day 15

until postpartum day 25, as reported in an abstract. The abstract did not state whether controls were

used. At 6 and 12 mg Hg/kg-day, pup weight was decreased on postpartum day 1. Subsequent weight

gain in these groups was also decreased. No other effects on development or behavior were observed

postpartum. Females at 6 and 12 mg Hg/kg-day had a decreased rate of weight gain, and gestation time

was slightly extended. Statistical analyses were not reported.

Pritchard et al. (1982b) administered 12, 16, or 24 mg mercuric chloride/kg-day (9, 12, 18 mg

Hg/kg-day) to female rats (strain and number not reported) by gavage before mating and during gestation. 

The abstract did not report whether controls were used. At 12 mg Hg/kg-day and above, females

exhibited weight loss and appeared unhealthy, estrous cycles became irregular, and high preimplantation

losses were observed. No effects on ovulation, estrous cycles, implantation, and fetal development were

observed at 9 mg Hg/kg-day. Statistical analyses were not reported.
3-54

Table 3-47

Developmental Toxicity of Inorganic Mercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/Long- Once 0, 2.0 Growth retardation; inhibition of eye formation in group Rizzo and Furst

Evans/5 F Gd 5, 12, or (HgO) treated on Gd 5, with some effect on Gd 12 group. 1972

19 Limitations: No statistical analysis; small number of

(gavage) litters in treated groups (and controls)

BML not reported

Rat/Strain 10 d 6, 9, 12, 18 Decreased fetal and placental weights (LOAEL = 9); McAnulty et al.

NS/no. F NS 1 x/d (HgCl ) malformations at 18. 1982

Gd 6-15 Limitations: Reported as an abstract; few details reported

(gavage) BML not reported

2

Rat/Strain Approx. 32 d 3, 6, 12 Decreased pup weight and weight gain (LOAEL = 6); no Pritchard et al.

NS/no. F NS 1 x/d (HgCl ) effect in an unspecified developmental and behavioral 1982a

Gd 15-ppd 25 testing battery.

(gavage) Limitations: Reported as an abstract; few details reported

2

BML not reported

Rat/Strain Before 9, 12, 18 High preimplantation loss (LOAEL = 12). Pritchard et al.

NS/no. F NS mating and (HgCl ) Limitations: Reported as an abstract; few details reported 1982b

during BML not reported

gestation

(gavage)

2

Hamster/ Once 0, 2.5, 5, 16, 22, Increased incidence of abnormal (small, retarded, Gale 1974

Golden/10 F Gd 8 32, 47, 63 edematous, and/or malformed--exencephaly,

exposed/3 F (gavage) [Hg (CH COO) ] encephalocele, ectrodactyly, etc.) fetuses (LOAEL = 16,

controls p<0.05); maternal toxicity: weight loss, diarrhea, slight 3 2

tremors, somnolence, tubular necrosis, hepatocellular

necrosis.

Limitation: Small sample size; smaller control group;

insufficient detail about number of animals sacrificed at

Gd 12 or Gd 14; single day of treatment; incomplete

examinations reported (no visceral, only partial skeletal)

BML not reported

In addition to the oral and inhalation studies summarized above, several studies using other

routes of administration (i.p., s.c., i.v.) provide evidence of developmental toxicity associated with

exposure to mercury salts. These studies are summarized below.

Gale and Ferm (1971) injected anesthetized pregnant golden hamsters (6–19/group)

intravenously with 0, 2, 3, or 4 mg mercuric acetate/kg (0, 1.3, 1.9, or 2.5 mg Hg/kg) on gestation day 8. 

Controls were injected with vehicle (demineralized water). Maternal animals were sacrificed on

gestation day 12 or 14, and the uterine contents were examined. A significantly increased incidence of

resorptions was observed at all doses. In addition, increased incidences of retarded and edematous

fetuses were observed at all doses (statistical significance not reported). Toxic effects observed in

maternal animals included weight loss, diarrhea, slight tremor, somnolence and kidney lesions; however,

the report did not specify at which doses the maternal effects were observed.
3-55

Gale (1974) compared the embryotoxicity of mercuric acetate administered by different routes in

pregnant golden hamsters (3–23/group). Subcutaneous administration of 0, 4, 8, 20, 35, or 50 mg

mercuric acetate/kg (0, 2.5, 5, 13, 22, or 32 mg Hg/kg) on gestation day 8 resulted in a significant

decrease in the percentage of normal embryos and a significant increase in the percentage of small

embryos at 2.5 mg Hg/kg. At 5 mg Hg/kg, significant increases in resorptions, abnormal, retarded,

edematous, and malformed fetuses were observed. Intraperitoneal administration of 0, 2, 4, or 8 mg

mercuric acetate (0, 1.3, 2.5, or 5 mg Hg/kg) on gestation day 8 resulted in significant increases in the

percentage of resorptions, abnormal, small, and edematous fetuses at 1.3 mg Hg/kg. Intravenous

administration of 0 or 4 mg mercuric acetate/kg (0 or 2.5 mg Hg/kg) on gestation day 8 resulted in

significant increases in resorptions, abnormal, small, retarded, edematous, and malformed fetuses at 2.5

mg Hg/kg. Comparison of the extent of the developmental toxicity demonstrated an effect of route of

administration: i.p. > i.v. > s.c. > oral.

Gale (1981) compared the embryotoxicity of mercuric acetate in 6 strains of hamsters

(LAK:LVG[SYR], CB/SsLak, LHC/Lak, LSH/SsLak, MHA/SsLak. PD4/Lak). Maternal animals of the

various strains (3–9/group) were injected subcutaneously with 0 or 15 mg/kg mercuric acetate (0 or 9.5

mg Hg/kg) on gestation day 8. Controls were injected with demineralized distilled water. Maternal

animals were sacrificed on gestation day 12 or 15, and the uterine contents were examined. The

percentage of resorptions was significantly increased in all strains examined on days 12 and 15. Strainspecific variations were observed in the incidences of abnormal, edematous and retarded fetuses and in

the incidences of ventral wall defects, distension of the pericardial cavity, cleft palate, hydrocephalus and

cardiac abnormalities. Maternal toxicity was not described.

Kavlock et al. (1993) injected pregnant Sprague-Dawley rats (6–25/group) subcutaneously with

0, 1, 2, 3, or 4 mg mercuric chloride/kg (0, 0.7, 1.5, 2.2, or 3.0 mg Hg/kg) on gestation day 7, 9, 11, or 13. 

On gestation day 21, rats were sacrificed, and the uterine contents were examined. No increase in

malformations was observed in fetuses from mercuric chloride-treated dams. Exposure on gestation day

7 resulted in a significant decrease in fetal weight and an increase in the number of supernumerary ribs

(statistical significance not reported) at 2.2 mg Hg/kg. Exposure on gestation day 9 resulted in

significantly decreased live fetuses/litter and increased resorptions at 3 mg Hg/kg. Exposure on gestation

days 11 or 13 resulted in no significant differences in fetal parameters. Maternal toxicity (increased

mortality, decreased body weight, increased kidney weight, increased urine osmolality, and/or increased

serum urea or creatinine) were observed at 1.5 mg Hg/kg and above. No consistent correlations were

observed between maternal and fetal toxicity.

Kajiwara and Inouye (1986) injected Kud:ddY mice (10/group) intravenously with 0, 0.5, 1.0,

1.5, 2.0, and 2.5 mg Hg/kg as mercuric chloride on gestation day 0. Controls were injected with vehicle

(physiological saline). Maternal animals were sacrificed on gestation day 3.5, and the oviducts and

uterus were flushed to obtain preimplantation embryos. At 1.5 mg Hg/kg and above, the number of

abnormal embryos was significantly increased. Maternal animals at 1.5 mg Hg/kg and above exhibited a

decrease in body weight (statistical significance not determined). The study authors suggested that fetal

toxicity may have been related to maternal toxicity.

Kajiwara and Inouye (1992) injected Kud:ddY mice (5–15/group) intravenously with 0, 1, 2, or

2.5 mg Hg/kg as mercuric chloride on gestation day 0. Controls were injected with vehicle

(physiological saline). Maternal animals were sacrificed on gestation day 5 or 12, and the oviducts and

uterine contents were examined. The animals sacrificed at gestation day 5 showed statistically

significant decreases in the number of embryos at all doses and an increase in blastocysts without

decidua (delay of implantation) at 2 and 2.5 mg Hg/kg. The animals sacrificed at gestation day 12

showed a statistically significant decrease in the number of implants, number of living fetuses and
3-56

average fetal weight at 2 and 2.5 mg Hg/kg. Maternal toxicity was not well-described, but 7 females at

2.5 mg Hg/kg and 2 females at 2 mg Hg/kg died. The study did not determine whether the failure to

implant was due to fetal toxicity or maternal uterine dysfunction.

A study by Bernard et al. (1992) was performed to assess whether prenatal and early postnatal

exposure to inorganic mercuric salts can produce nephrotoxic effects. They found that s.c. injection of

dams with just 1 mg/kg-day during pregnancy caused renal effects in the offspring. Of note is that these

effects did not appear to be significantly different in the group of dams dosed throughout the gestation

period compared to dams dosed only during the last 8 days of gestation.

3.2.3.9 Reproductive

A single case study was located concerning reproductive toxicity in humans exposed to inorganic

mercury; however, it is not clear whether the effects were compound-related. No information was

identified regarding the reproductive toxicity of inorganic mercury following inhalation exposure.

Table 3-48

Reproductive Toxicity of Inorganic Mercury in Humans: Case Study

Species/ Exposure Dose

No. per Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/1 F Once 30 Spontaneous abortion 13 days after ingestion of mercuric Afonso and

(tablet) (HgCl ) chloride deAlvarez 1960 2

Limitation: Case study; abortion may have been unrelated

to mercury exposure

BML not reported

In animals orally exposed to inorganic mercury compounds, changes in the estrous cycle and

ovulation and/or increased resorptions were reported (Pritchard et al. 1982b).

In male mice administered a single i.p. dose of 0.74 mg Hg/kg as mercuric chloride, fertility

decreased between days 28 and 49 post-treatment with no obvious histological effects noted in the sperm

(Lee and Dixon 1975). The period of decreased fertility indicated that spermatogonia and premeiotic

spermatocytes were affected. The effects were less severe than those noted after treatment with a similar

dose of methylmercury. A single i.p. dose of 1.5 mg Hg/kg as mercuric chloride administered 1–5 days

prior to mating in female mice resulted in a significant decrease in the total number of implants, number

of living embryos and a significant increase in the percentage of dead implants (Suter 1975). These

effects suggest that mercury may be a weak inducer of dominant lethal mutations. In female golden

hamsters administered 6.4 or 12.8 mg Hg/kg subcutaneously, there was no observed increase in

chromosomal aberrations in metaphase II oocytes (Watanabe et al. 1982). At the first estrous cycle post

treatment, there was a significant increase in the number of degenerated oocytes in animals at the highdose group. At the second estrous cycle both treatment groups had increased numbers of degenerated

oocytes, suggesting an effect of mercuric chloride on ovulation.
3-57

Table 3-49

Reproductive Toxicity of Inorganic Mercury in Animals; Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/strain NS "before 9, 12, 18 Irregular estrous cycles and high preimplantation loss Pritchard et al.

NS/NS no. of F mating and (HgCl ) (LOAEL = 12); decreased ovulation (LOAEL = 18) 1982b

during Limitations: Limited details; reported as an abstract; no

gestation" statistical analysis reported

(gavage) BML not reported

2

3.2.3.10 Genotoxicity

Two occupational studies (Anwar and Gabal 1991; Popescu et al. 1979) reported on workers

inhaling inorganic mercury; the data were inconclusive regarding the clastogenic activity of inorganic

mercury. Workers involved in the manufacture of mercury fulminate (Hg[OCN] ) had a significant 2

increase in the incidence of chromosomal aberrations and micronuclei in peripheral lymphocytes when

compared to unexposed controls (Anwar and Gabal 1991). There was no correlation between urinary

mercury levels or duration of exposure to the increased frequency of effects; the study authors concluded

that mercury may not have been the clastogen in the manufacturing process. In a study by Popescu et al.

(1979), 18 workers exposed to a mixture of mercuric chloride, methylmercuric chloride and

ethylmercuric chloride had significant increases in the frequency of acentric fragments (chromosome

breaks). The findings, however, are suspect because the control group was not matched for sex, smoking

habits or sample size.

Table 3-50

Genotoxicity of Inorganic Mercury in Humans

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/29 M 20.8 yr (avg) NS Increased incidence of chromosomal aberrations (p<0.001) Anwar and Gabal

exposed/ 29 M (occup) Hg(OCN) and micronuclei (p<0.01) in lymphocytes of workers exposed 1991

control to mercury fulminate compared with age-matched controls; no 2

correlation between frequency of chromosome and exposure

duration or urinary mercury level.

BML: 123.2 μg/L in urine (avg)

Human/18 M 10.5 yr 0.15-0.44 Increased frequency of chromosomal breaks. Popescu et al. 1979

exposed/ (occup) (HgCl ) Limitations: Workers also exposed to methylmercuric

10 control chloride and ethylmercuric chloride, and one worker had 2

history of benzene poisoning; control group was not matched

for sex, smoking habits, or sample size.

BML: 890 μg/L in urine (avg)
3-58

Exposure to inorganic mercury may produce an increase in chromosomal aberrations in mice

following oral and inhalation exposure (Ghosh et al. 1991; Selypes et al. 1984). Mercuric chloride

administered to mice by gavage induced a dose-related increase in chromosome aberrations and aberrant

cells in the bone marrow (Ghosh et al. 1991); however, mice given i.p. doses of mercuric chloride have

shown no increase in chromosomal aberrations in bone marrow cells (Poma et al. 1981) and no increase

in aneuploidy in spermatogonia (Jagiello and Lin 1973). Similarly, an increased incidence of

chromosomal aberrations (primarily deletion and numeric aberrations) was observed in livers of fetal

mice exposed to mercury in utero as the result of maternal inhalation of aerosols of mercuric chloride

(Selypes et al. 1984). Female golden hamsters injected s.c. with mercuric chloride were observed to have

increased incidence of chromosome aberrations in bone marrow cells but not in metaphase II oocytes

(Watanabe et al. 1982). Mercuric chloride concentrations in the ovaries were low but had an inhibiting

effect on ovulation. Verschaeve et al. (1984) reported that in vitro exposure of human lymphocytes and

muntjac fibroblasts to mercuric chloride resulted in segregation abnormalities; namely, c-mitotic figures. 

The effects of mercuric chloride on genetic material has been suggested to be due to the ability of

mercury to inhibit of the formation of the mitotic spindle, which can result in c-mitotic figures. Mercuric

chloride has also been shown to inhibit nucleolus organizing activity in human lymphocytes (Vershaeve

et al. 1983).

Positive dominant lethal results have been obtained in studies in which rats were administered

mercuric chloride orally (Zasukhina et al. 1983). Suter (1975) observed a small, but significant increase

in the number of non-viable implants when female mice were administered mercuric chloride by

intraperitoneal injection; this effect was not observed when males were treated. It was not clear whether

the increase in non-viable implants was due to maternal toxicity or to a true dominant lethal effect of the

treatment.

Sex-linked recessive lethal mutations were not observed as a consequence of exposure of male

Drosophila melanogaster by either feeding or injection (NTP 1993).

As summarized in NTP (1993) and U.S. EPA (1985), mercuric chloride has produced some

positive results for clastogenicity in a variety of in vitro and in vivo genotoxicity assays, but mixed

results regarding its mutagenic activity have been reported. Mercuric chloride was negative in gene

mutation tests with Salmonella typhimurium strains TA1535, TA1537, TA98 and TA102 with or without

hepatic microsomal preparations (S9) (Arlauskas et al. 1985; Marzin and Phi 1985; Wong 1988). 

Mercuric chloride has shown evidence of DNA damage in the Bacillus subtilis rec assay (Kanematsu et

al. 1980) but did not induce lytic phage in a lysogenic E. coli strain (Rossman et al. 1984).

DNA damage (single strand breaks) has also been observed in assays using rat and mouse

embryo fibroblasts (Zasukhina et al. 1983) and Chinese hamster ovary (CHO) cells and human KB cells

(Cantoni and Costa 1983; Cantoni et al. 1982, 1984a,b; Christie et al. 1984, 1986; Robison et al. 1982,

1984; Williams et al. 1987). Mercuric chloride also produced chromosome aberrations and sister

chromatid exchange (SCE) in CHO cells (Howard et al. 1991) and SCE in human leucocytes (Morimoto

et al. 1982). Negative results for chromosomal aberrations were reported for FM3A cells (from a mouse

mammary carcinoma) (Umeda and Nishimura 1979) and for two human diploid lines, WI38 and MRC 5

(Paton and Allison 1972). Negative results for SCE were reported for don cells (Ohno et al. 1982) and

for P388D, mouse cells and CHO cells (Anderson 1983). Evidence of gene mutations (considered

weakly positive) was observed in L5178Y mouse lymphoma cells in the presence of microsomal

preparations (Oberly et al. 1982).

NTP (1993) reached the following conclusions from their in vitro testing of mercuric chloride:

not mutagenic for Salmonella typhimurium in preincubation protocols with and without rat and hamster
3-59

liver preparations; positive for L5178Y cells without addition of hepatic preparations; negative for SCE

in CHO cells without addition of S9 but weakly positive when rat S9 was added; positive for

chromosomal aberrations in CHO cells in the absence but not the presence of liver preparations; it was

not clear what role was played by cytotoxicity in the generation of these chromosomal aberrations.

Table 3-51

Genotoxicity of Inorganic Mercury in Animals

 Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/m ) Effects/Limitations/BML Reference 3

Mice/CFLP 4 d 0.17, 1.6 Increased incidence of chromosomal aberrations in fetal Selypes et al. 1984

NS/No. F NS 4 hr/d (HgCl ) hepatocytes

Gd 9-12 Limitations: The number of mothers corresponding to the 10 2

fetuses examined was not reported; no statistical analysis

BML not reported

3.3 Methylmercury

Organic mercury compounds have been used as fungicides and as pharmaceutical agents

(diuretics). Organic mercurials including Metaphen, Merthiolate and Mercurochrome still find use as

topical antiseptics. Phenylmercury salts are used in pharmaceutical, ophthalmic and cosmetic

preparations to control growth of microbial organisms (Joklik et al. 1984). Other organic mercury

compounds include methylmercuric chloride (MMC), methylmercuric hydroxide (MMH) and

phenylmercuric acetate (PMA). Nearly all of the available toxicity studies for organic mercury

compounds, however, are for methylmercury. Unless otherwise noted, all studies summarized in tables

in this section are for methylmercury. All oral doses were converted to mg Hg/kg-day, and all inhalation

doses were converted mg Hg/m using the method shown in Appendix A. 3

3.3.1 Critical Noncancer Data

This section provides descriptions of studies considered by U.S. EPA in evaluation of systemic

health endpoints, largely neurotoxicity in exposed adults and in children exposed in utero. Chapter 6

describes the derivation of an RfD for methylmercury based on developmental neurologic abnormalities

in human infants. For completeness some of these studies are also presented in subsequent sections in

tabular form.

3.3.1.1 Human Data

Several studies of methylmercury poisonings in humans have been reported (see discussion on

Neurologic Effects). CNS effects were observed in several studies summarized by Clarkson et al. (1976),

Nordberg and Strangert (1976), and WHO (1976). CNS effects including ataxia and paresthesia have

been observed in subjects with blood mercury concentrations as low as 200 μg/L. 
3-60

The original epidemiologic report of methylmercury poisoning involved 628 human cases that

occurred in Minamata, Japan, between 1953 and 1960. The overall prevalence rate for the Minamata

region for neurologic and mental disorders was 59%. Among this group 78 deaths occurred, and hair

concentrations of mercury ranged from 50–700 ppm. These hair mercury concentrations were

determined through the use of less precise analytic methods than were available for later studies. The

most common clinical signs observed in adults were paresthesia, ataxia, sensory disturbances, tremors,

impairment of hearing and difficulty in walking. Examination of the brains of severely affected patients

that died revealed marked atrophy of the brain (55% normal volume and weight) with cystic cavities and

spongy foci. Microscopically, entire regions were devoid of neurons, granular cells in the cerebellum,

golgi cells and Purkinje cells. Extensive investigations of congenital Minamata disease were undertaken,

and 20 cases that occurred over a 4-year period were documented. In all instances the congenital cases

showed a higher incidence of symptoms than did the cases wherein exposure occurred as an adult. 

Severe disturbances of nervous function were described, and the affected offspring were very late in

reaching developmental milestones. Hair concentrations of mercury in affected infants ranged from 10 to

100 ppm. Hair mercury levels for the mothers during gestation were not available (Tsubaki 1977).

In 1971, an unknown number of people in Iraq were exposed to methylmercury-treated seed

grain that was used in home-baked bread. Studies conducted on this population include that of Bakir et

al. (1973), Marsh et al. (1987) and others. Toxicity was observed in many adults and children who had

consumed this bread over a three-month period, but the population that showed greatest sensitivity were

offspring of pregnant women who ate contaminated bread during gestation. The predominant symptom

noted in adults was paresthesia, and it usually occurred after a latent period of from 16 to 38 days. In

adults symptoms were dose-dependent, and among the more severely affected individuals ataxia, blurred

vision, slurred speech and hearing difficulties were observed. Signs noted in the infants exposed during

fetal development included cerebral palsy, altered muscle tone and deep tendon reflexes, as well as

delayed developmental milestones (e.g., walking before 18 months and talking before 24 months). Some

information indicated that male offspring were more sensitive than females. The mothers experienced

paresthesia and other sensory disturbances but at higher doses than those associated with their children

exposed in utero. Unique analytic features of mercury (analysis of segments of hair correlated to specific

time periods in the past) permitted approximation of maternal blood levels that fetuses were exposed to

in utero. The data collected by Marsh et al. (1987) summarized clinical neurologic signs of 81 mother

and child pairs. From x-ray fluorescent spectrometric analysis of selected regions of maternal scalp hair,

concentrations ranging from 1 to 674 ppm were determined, then correlated with clinical signs observed 

in the affected members of the 81 mother-child pairs. Among the exposed population there were affected

and unaffected individuals throughout the dose-exposure range. (See also Bakir et al. (1973) in sections

on Death and Neurological Effects).

McKeown-Eyssen et al. (1983) provided a report of neurologic abnormalities in four

communities of Cree Indians in northern Quebec. A group of 247 children between 12 and 30 months of

age was evaluated for clinical signs consistent with methylmercury exposure. A pediatric neurologist

evaluated the children for the following: height, weight, head circumference, dysmorphic and congenital

features and the presence of acquired disease. In addition to the DDST the following assessments were

done: special senses, cranial nerve function, sensory function, muscle tone, stretch reflexes, coordination, persistence of Babinski response, and a summary of signs for absence or presence of

neurologic abnormality. An attempt was made to account for possible confounding factors; the

interviewers determined alcohol and tobacco consumption patterns among the mothers of affected

children. Age of the mothers and multiparity was also taken into account in analysis of the data.

Maternal hair mercury was used as the exposure measure. The average maternal hair mercury was the 

same for boys and girls (6 ppm); only 6% of the population had exposure above 20 ppm. The prevalence

of multiple abnormal neurologic findings was about 4% for children of both sexes. The most frequently
3-61

observed abnormality was delayed deep tendon reflexes. This was seen in 11.4% of the boys and 12.2%

of the girls. Abnormality of muscle tone or reflexes showed a significant positive association with

maternal mercury exposure for boys, but not for girls. A consistent dose-response relationship for this

effect was not observed; however, the greatest prevalence of the effect in boys occurred for those with

mothers in the highest exposure group (13.0-23.9 ppm mercury in hair). No other measure of abnormal

or decreased neurologic function or development showed a significant positive association with maternal

hair mercury. The prevalence of abnormality of muscle tone or reflexes was found to increase 7 times

with each increase of 10 ppm of the prenatal exposure index. There was possible influence of alcohol

consumption and smoking among mothers on the effects observed in their children.

Studies performed in New Zealand investigated the development of children who had prenatal

exposure to methylmercury (Kjellstrom et al. 1986a, 1989). A group of 11,000 mothers who regularly

ate fish was initially screened by survey; of these, about 1000 had consumed 3 fish meals per week

during pregnancy. Working from these two large groups, 31 matched pairs were established. A

reference child matched for mother's ethnic group, age, child's birthplace and birth date was located for

each child in the high fish consumption group. Mercury exposure during gestation was determined from

maternal hair analysis. The average hair concentration for high exposure mothers was 8.8 ppm and for

the reference group was 1.9 ppm. At 4 years of age, the children were tested using the Denver

Developmental Screen Test (DDST). This is a standardized test of a child's mental development and can

be administered in the child's home. It consists of four major function sectors: gross motor, fine motor,

language and personal-social. A developmental delay in an individual item is scored when the child has

failed in his/her response and at least 90% of children can pass this item at a younger age. The whole test

is scored as abnormal, questionable, or normal. Standardized vision tests and sensory tests were also

performed to measure development of these components of the nervous system. The prevalence for

developmental delay in children was 52% for progeny of high mercury mothers and 17% for progeny of

mothers of the reference group. The hair mercury concentrations of the mothers in this study were lower

than those associated with CNS effects in children exposed in Japan and Iraq. The results of the DDST

included 2 abnormal scores and 14 questionable scores in the high mercury-exposed group and 1

abnormal and 4 questionable scores in the control group. The results remained statistically significant

after the 8 pairs where ethnic group matching was not successful and twins were excluded. Analysis of

the DDST results by sector showed that developmental delays were most commonly noted in the fine

motor and language sectors, but the differences between the experimental and control groups were not

significant. The differences noted in performance of the DDST between high mercury-exposed and

referent children could be due to confounding variables. Infants of the mercury-exposed group more

frequently had low birth weights and were more likely to be born prematurely. 

A second stage follow-up of the original Kjellstrom study was carried out when the children

were 6 years old (Kjellstrom et al. 1986b). In this later study the high exposure children were compared

with 3 control groups with lower prenatal mercury exposure. During pregnancy, mothers in two of these

control groups had high fish consumption and average hair mercury concentrations of 3–6 ppm and

0–3ppm, respectively. The high exposure group was matched with controls for maternal ethnic group,

age, smoking habits, residence, and sex of the child. For the second study, 61 of 74 high exposure

children were available for study. Each child was tested with an array of scholastic, psychological and

behavioral tests which included Test of Language Development (TOLD), the Wechsler Intelligence Scale

for Children and McCarthy Scale of Children's Abilities. The results of the tests were compared between

groups. Confounding was controlled for by a modelling procedure using linear multiple regression

analysis. A principal finding was that normal results of the psychological test variables were influenced

by ethnic background and social class. High prenatal methylmercury exposure decreased performance in

the tests, but it contributed only a small part of the variation in test results. It was found that an average

hair mercury level of 13–15 ppm during pregnancy was consistently associated with decreased test
3-62

performance. Size of the experimental groups limited the power of the study to determine if lower

exposure levels might have had a significant effect on test results. The studies are limited for assessing

methylmercury toxicity because the intelligence tests used may not be the most appropriate for defining

the effects of methylmercury. Also, greater significance was seen in differences of cultural origins of the

children than the differences in maternal hair methylmercury concentrations.

A prospective study (Marsh et al. 1995) of fetal exposure to methylmercury through maternal

consumption of fish was conducted in the Peruvian fishing village of Mancora between 1981-1984. An

account of the study was written in 1985 and subsequently published in 1995. Hair samples were

obtained from 369 pregnant women, and neurologic examinations were performed on 194 of the children

from these pregnancies. Of this cohort there were 131 mother-infant pairs with complete clinical data

and hair samples that provided quantitative measures of methylmercury exposure during pregnancy. The

peak hair mercury levels ranged from 1.2 to 30 ppm with a mean of 8.3 ppm. Testing of hair segments

and statistical analyses confirmed that the women had reached a relatively steady state in terms of

methylmercury. No significant correlation was shown between increasing maternal hair methylmercury

and effects on the developmental determinations assessed in this study. Measures included the

following: perinatal factors (labor or delivery difficulty, abnormal respiration or color), maternal

paresthesia, speech retardation, muscle tone evaluation, determination of primitive and tendon reflexes,

ataxia and mental and motor retardation.

There has been an increasing concern about the likelihood of exposure of people residing in the

Amazonian watershed to methylmercury in fish and other sources. The Amazon River valley is the site

of many small gold mining operations which use metallic mercury in the extraction process; it is

estimated that 55-60% is released to the atmosphere and 40-45% enters the aquatic environment (Pfeiffer

et al. 1989; Malm et al 1990). Lebel et al (1996) have published a study of 29 adult residents of two

villages located on the Tapajos River, a tributary of the Amazon located about 200 Km from the mining

sites. There were 14 women and 15 men aged 35 and younger who were randomly chosen from a

previous survey. Total hair mercury ranged from 5.6 to 38.4 ppm; methylmercury constituted between

72.2% and 93.3% of the sampled mercury. A quantitative behavioral neurophysiologic battery was

modified for administration to persons with minimal formal education in an area without electricity. For

women only there was a decrease with increasing hair mercury in manual dexterity as measured in the

Santa Ana test, Helsinki version. For both men and women there was a statistically significant decrease

with increasing mercury in color discrimination capacity (as measured by the Lanthony D-15 desaturated

panel). Near visual contrast sensitivity profiles (measured with the Vistech 6000) and peripheral visual

field profile (Goldman Perimetry with Targets I and IV) were both reduced in the individuals with the

highest hair mercuries. The authors noted that constriction of the visual field has been observed in other

instances of mercury intoxication and that changes in contrast sensitivity has been noted in non-human

primates exposed to methylmercury (Rice and Gilbert 1982; Rice and Gilbert 1990).

Lonky et al. (1996) have reported on studies of behavioral effects in newborns as a consequence

of maternal consumption of fish from Lake Ontario in the U.S.. Fish from the U.S. Great Lakes have

been shown to be contaminated with a number of environmental pollutants including PCBs and

methylmercury. A total of 559 children were tested. Exposure was measured as fish consumption which

was determined by interviews taking place during pregnancy. Information was collected on fish species,

number of meals, serving size, and method of preparation. Exposure was scaled in PCB equivalent

weights. Women were assigned to the high dose group if they reported eating at least 40 PCB-equivalent

pounds of fish in their lifetime (n=152); the low fish consumption group numbered 243, and there were

164 no-fish-consumption controls. Infants born to the high fish consumption group scored significantly

more poorly on the Reflex, Autonomic and Habituation clusters of the Neonatal Behavioral Assessment

Scale. The authors do not attribute these developmental effects to exposure to any one compound;
3-63

analyses of cord blood and maternal hair are being done for PCBs, DDE, hexachlorobenzene, lead and

mercury. 

In 1981 a group of researchers in conjunction with the Seychelles Island government initiated a

large study on developmental effects of low level methylmercury exposure from consumption of marine

fish. The Seychelles is an island country in the Indian Ocean near the coast of Africa where fish are

consumed by the population on a daily basis. A pilot study (essentially a cross-sectional study) was

initiated which focused on all children born between February 1989 and February 1990. A total of 804

mother-infant pairs were enrolled, which was 48% of those eligible. The main study was designed to be

prospective and involved 779 mother-child pairs. In both studies maternal hair samples and umbilical

cord blood were measured for total mercury content using atomic absorption spectroscopy. Children

were enrolled in the pilot study during the years 1987-1988; the main study enrolled children from 1989

to 1990. Both the pilot and main studies involved about 50% of all children born during the year, 804

and 779 children respectively. (Shamlaye et al. 1995)

The authors have noted that the pilot study was not as well-controlled as the main or longitudinal

study: there were fewer covariates, medical records were not reviewed as carefully, there was less

information on socio-economic status. Subsets of enrolled children were tested, and the main purpose of

testing was to pilot the test batteries. Multiple hair samples collected for the pilot study during gestation

had methylmercury ranging from 0.6 to 36.4 ppm with a median of 6.6 ppm (Myers et al. 1995b). The

endpoints evaluated during the pilot study included a general neurologic evaluation and the DDST-R in

addition to examinations of physical development. The age of each child was known; children were

tested once between the ages of 5 and 109 weeks of age. In addition to DDST-R testing, a medical

history was obtained, and general and neurologic examinations were also performed. Statistical analysis

included the following covariates: gender, birth weight, one- and five-minute Apgar score, age at testing,

and medical problems. Covariates for the mothers were age, tobacco and alcohol consumption during

pregnancy and medical problems. An association between in utero mercury exposure was found for

DDST-R abnormal plus questionable scores combined. (Myers et al 1995b)

A subset of the pilot cohort was evaluated at 66 months (Myers et al. 1995a). This group of 217

children was administered the McCarthy Scales of Children’s Abilities, the Preschool Language Scale,

and the Letter-Word Recognition and Applied Problems subtests of the Woodcock-Johnson Tests of

Achievement that were appropriate to the children’s age. The median maternal hair mercury for this

group was 7.11 ppm. Mercury exposure (measured as maternal hair mercury) was negatively associated

with four endpoints: The McCarthy General Cognitive Index and Perceptual Performance subscale; and

the Preschool Language Scale Total Language and Auditory Comprehension subscale. When statistically

determined outliers and points considered to be influential were removed from the analyses, statistical

significance of the association remained only for auditory comprehension. 

The prospective or main study, involved evaluation of children at 6.5, 19, 26 and 66 months of

age. Age-appropriate tests administered included the following: Infantest (or Fagan’s test of visual

recognition memory), Bayley Scales of Infant Development (BSID), McCarthy Scales of Children’s

Abilities, the Preschool Language Scale and the DDST (6.5 months only). Maternal intelligence and

home environment were also assessed (Marsh et al. 1995a). In the group (n = 740) evaluated at 6.5

months, median maternal hair mercury was 5.9 ppm with a range of 0.5 ppm to 26.7 ppm. No association

with maternal hair mercury was found for any of six endpoints in six children tested at six months.

(Myers et al. 1995c).

Evaluations at 19 and 29 months were done on groups of 738 and 736 individuals, respectively

(Davidson et al. 1995). Median maternal hair mercury was 5.9 ppm and the range was 0.5 to 26.7 ppm.
3-64

Children were evaluated with the BSID at 19 months of age. At 29 months, children were administered

the BSID as well as the Bayley Infant Behavior Record. Mean BSID mental indices were comparable to

scores for children in the U.S. at both 19 and 29 months. The BSID Psychomotor Scale was reflective of

the rapid development of motor skills by children reared in African cultures. No effects of mercury

exposure were seen on outcome of five test endpoints at 19 months. At 29 months there was an

association between mercury exposure and decreased activity level in male children only. The authors

point out that the activity level observed during the testing session may not reflect the child’s activity

level in other settings. There was no association with maternal hair mercury in 15 other endpoints.

Delayed onset of walking and talking were among the measures of toxicity used in the evaluation

of Iraqi children in utero (Marsh et al. 1987.) When the children in the Seychelles study reached the age

of 19 months, caregivers were queried as to the time at which they walked and talked (n=760 and 680,

respectively). Onset of walking was defined as walking without support and age at talking as the age at

which the child said words other than “mama” or “dada.” The mean age for walking was 10.7 months for

girls and 10.6 for boys; for talking it was 10.5 for girls and 11.0 for boys. For female children, the mean

age for walking was similar across all mercury exposures. In boys, the mean age for walking increased

between 0.3 and 0.6 months from the lowest to highest mercury exposure groups. Statistical analyses

adjusted covariates and outlying points data. With these adjustments there was no statistically significant

association between maternal hair mercury and age at which the child walked or talked.

The overall conclusion of the studies published to date is that it is yet unclear whether an

association exists between low level mercury exposure and neurologic deficits in children. The study

does show a close correlation between maternal hair mercury and neonatal levels of mercury in brain

tissue (Cernichiari et al. 1995). The authors cautioned in several papers that subtle neurologic and

neurobehavioral effects are more likely to be detected in older rather than younger children. The overall

conclusion of the authors is that their results require careful interpretation, and that an association

between relatively low level mercury exposure in utero and neurologic deficits has not been conclusively

demonstrated. 

A large study was initiated in the Faroe Islands in 1986 on neurologic developmental effects of

methylmercury and PCB exposure in utero (Grandjean et al. 1997). The population of the Faroes is

relatively homogeneous. During pregnancy consumption of alcoholic beverages is uncommon. As with

other fishing communities, seafood is large part of the Faroese diet. Increased mercury exposure,

however, is largely attributed to the eating of pilot whale, which is traditionally hunted and shared among

the population (Grandjean et al. 1992a). Subjects were a group of 917 (of an initial cohort of 1022)

children born between 1986 and 1987 and evaluated at about 7 years of age. Mercury was measured in

maternal hair and cord blood, and a subset of cords was evaluated for PCBs. Of the initial cohort the

median blood mercury was 24.2 μg/L with 25% of the samples above 40 μg/L. The median maternal hair

mercury concentration was 4.5 ppm, and 13% were greater than 10 ppm (Grandjean et al. 1992a). Cord

blood mercury was found to be most closely associated with maternal hair mercury; the association with

hair measurements in the children at 12 months and 7 years was not as strong (Grandjean et al. 1997).

 At seven years children received a physical examination including a functional neurological

examination which emphasized motor co-ordination and perceptual-motor performance. Visual acuity

was measured using Snellen’s board; contrast sensitivity was assessed using the Functional Acuity

Contrast Test. Standard hearing tests were done. Neurophysiological tests included the following:

pattern-reversal visual evoked potentials; brainstem auditory evoked potentials; postural sway under four

conditions; and coefficient of variation for R-R intervals on electrocardiogram as a measure of autonomic

nervous system function. Neuropsychological tests included these: motor tests--the Neurobehavioral
3-65

Evaluation System (NES) finger tapping test and the NES Hand-Eye Coordination test; tactile processing

-- Tactual Performance Test; vigilance/attention--NES Continuous Performance Test; attention and

tracking -- Wechsler Intelligence Scale for Children-revised (WISC-r) digit span forward; reasoning and

cognitive flexibility--WISC-r similarities; visuospatial--WISC-r Block Designs and Bender Visual Motor

Gestalt Test; language--Boston Naming Test; memory--California Verbal Learning Test; and mood--

Nonverbal Analogue Profile of Mood States. 

Three neurological tests were found to be difficult for 7-year-old children; fewer than 60% of the

children performed optimally. On one of these tests, finger opposition, the group of 465 with optimal

performance had a geometric mean cord blood mercury of 21.8 μg/L of 21.8 by contrast to 23.9 μg/L for

the 425 children with questionable or deficient performance; this was a statistically significant

difference. The neurophysiological tests showed no indication of mercury-associated dysfunction.

Significant negative associations were seen on several neuropsychological tests. Even with inclusion of

covariates with uncertain influence on these tests results, multiple regression analysis indicated that 9/20

measures showed mercury related decrements (p< 0.05, one tailed). Application of a Peters-Belson

adjustment resulted in significant mercury associations for 11/20 measures. 

Pilot whale fat (blubber) is consumed in the Faroe Islands, and this could result in increased

exposure to PCBs. PCB concentrations in Faroese breast milk has been shown to be higher than in other

Scandinavian countries (Grandjean et al. 1995b). PCB determinations were done on a total of 436 cord

bloods, and PCB exposure was included as a covariate in the regression analyses. This had an effect only

on the regression for the Boston Naming Test. The authors concluded that in utero exposure to

methylmercury affects several domains of cerebral function. After exclusion of children with maternal

hair mercury concentrations above 10 ppm, the association between mercury exposure and

neuropsychological dysfunction remained unchanged. The authors, therefore, concluded that adverse

effects are observed at exposures below 10 ppm maternal hair (Grandjean et al. 1997). 

3.3.1.2 Animal Data

Rice (1989b) dosed five cynomolgus monkeys (Macaca fascicularis) from birth to 7 years of age

with 0.05 mg Hg/kg-day as methylmercuric chloride and performed clinical and neurologic examinations

during the dosing period and for an additional 6 years. As a sensitive indicator of the latent effects of

methylmercury, neurologic examinations performed at the end of the observation period revealed

insensitivity to touch and loss of tactile response. In the later stages of the observation period monkeys

dosed with methylmercury were clumsier and slower to react when placed in the exercise cage than were

unexposed monkeys.

Gunderson et al. (1986) administered daily doses of 0.04–0.06 mg Hg/kg as methylmercuric

hydroxide to 11 crab-eating macaques (Macaca fascicularis) throughout pregnancy, resulting in maternal

blood levels of 1080–1330 g/L in mothers and 1410–1840 g/L in the offspring. When tested 35 days

after birth, the infants exhibited visual recognition deficits.

Groups of 7 or 8 female crab-eating macaques (Macaca fascicularis) were dosed with 0.05 or

0.09 mg/kg-day of methylmercury through 4 menstrual cycles (Burbacher et al. 1984). They were mated

with untreated males, and clinical observations were made for an additional 4 months. Two of 7 highdose females aborted, and 3 did not conceive during the 4-month mating period. The other two females

delivered live infants. Two of 7 females exposed to 0.05 mg/kg-day aborted; the remaining 5 females

delivered live infants. All control females conceived, and 6 delivered live infants. These reproductive

results approached but did not reach statistical significance. Reproductive failure within dose groups

could be predicted by blood mercury levels. The dams did not show clinical signs of methylmercury
3-66

poisoning during the breeding period or gestation, but when females were dosed with 0.09 mg/kg-day for

a year, 4 of 7 did show adverse neurologic signs.

Bornhausen et al. (1980) has reported a decrease in operant behavior performance in 4-month-old

rats whose dams had received methylmercuric chloride on gestation days 6–9. A statistically significant

effect was seen in offspring whose dams had received 0.01 and 0.05 mg/kg five times during gestation. 

The authors postulated that more severe effects of in utero exposure would be seen in humans since the

biological half-time of mercury in the brain of humans is 5 times longer than the rat. In addition, much

longer in utero exposure to mercury would occur in humans since gestation is much longer.

Groups of Wistar rats (50/sex/group) were administered daily doses of 0.002, 0.01, 0.05, and

0.25 mg Hg/kg-day as methylmercuric chloride for 26 months (Munro et al. 1980). Female rats that

received 0.25 mg/kg-day had reduced body weight gains and showed only minimal clinical signs of

neurotoxicity; however, male rats that received this dose did show overt clinical signs of neurotoxicity,

had decreased hemoglobin and hematocrit values, had reduced weight gains and showed increased

mortality. Histopathologic examination of rats of both sexes receiving 0.25 mg/kg-day revealed

demyelination of dorsal nerve roots and peripheral nerves. Males showed severe kidney damage, and

females had minimal renal damage. This study showed a NOAEL of 0.05 mg/kg-day and a LOAEL of

0.25 mg/kg-day.

A 2-year feeding study of methylmercuric chloride was conducted in B6C3F1 mice

(60 mice/sex/group) at doses of 0, 0.4, 2, and 10 ppm (0, 0.03, 0.15, and 0.73 mg Hg/kg-day in males; 0,

0.02, 0.11, and 0.6 mg Hg/kg-day in females) to determine chronic toxicity and possible carcinogenic

effects (Mitsumori et al. 1990). Mice were examined clinically during the study, and neurotoxic signs

characterized by posterior paralysis were observed in 33 males after 59 weeks and 3 females after 80

weeks in the 0.6-mg Hg/kg-day group. A marked increase in mortality and a significant decrease in body

weight gain were also observed in the high-dose males, beginning at 60 weeks. Post-mortem

examination revealed toxic encephalopathy consisting of neuronal necrosis of the brain and toxic

peripheral sensory neuropathy in both sexes of the high-dose group. An increased incidence of chronic

nephropathy was observed in the 0.11- and 0.6-mg Hg/kg-day males. These results indicated that

B6C3F1 mice are more sensitive to the neurotoxic effects of methylmercury than ICR mice.

Ultrastructural renal changes were also observed in rhesus monkeys treated with

0.08–0.12 mg/kg of methylmercury although clinical changes were not observed (Chen et al. 1983).

3.3.2 Cancer Data

3.3.2.1 Human Data

Three studies were identified that examined the relationship between methylmercury exposure

and cancer. No persuasive evidence of increased carcinogenicity attributable to methylmercury exposure

was observed in any of the studies. Interpretation of these studies, however, was limited by poor study

design and incomplete descriptions of methodology and/or results.

Tamashiro et al. (1984) evaluated the causes of death in 334 subjects from the Kumamoto

Prefecture who had been diagnosed with Minamata disease and died between 1970 and 1980. Minamata

disease was used as a surrogate for methylmercury exposure. The cases were fishermen and their

families who had been diagnosed with methylmercury poisoning (Minamata disease); thus, Minamata

disease was used as a surrogate for methylmercury exposure. The controls were selected from all deaths

that had occurred in the same city or town as had the cases and were matched on the basis of sex, age at
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death (within 3 years) and year of death; two controls were matched to each case. Malignant neoplasms

were designated as the underlying cause of death in 14.7% (49/334) of the cases and 20.1% (134/668) of

the controls. For 47 cases in which Minamata disease was listed as the underlying cause of death, the

investigators reanalyzed the mortality data and selected one of the secondary causes to be the underlying

cause of death in order to allow examination of the cases and controls under similar conditions and

parameters. The three cases for which Minamata disease was listed as the only cause of death were

excluded from further analysis. Using the Mantel-Haenzel method to estimate odds ratios, no significant

differences were observed between the cases and controls with respect to the proportion of deaths due to

malignant neoplasms among males, females, or both sexes combined. The estimated odds ratios and 95%

confidence intervals were 0.84 (0.49–1.43), 0.58 (0.28–1.21), and 0.75 (0.50–1.11) for males, females,

and both sexes combined. Similarly, no increases were observed among the cases relative to the controls

when malignant neoplasms were identified as a secondary cause of death or were listed on death

certificates as one of multiple causes of death. These data suggest that cancer incidence is not increased

in persons with overt signs of methylmercury poisoning when compared to persons for whom no

diagnosis of methylmercury poisoning had been made. Interpretation is limited, however, by potential

bias in designating the cause of death among patients with known Minamata disease and by the

uncertainty regarding the extent of methylmercury exposure and undiagnosed Minamata disease among

the controls.

In a subsequent study, Tamashiro et al. (1986) compared the mortality patterns (between 1970

and 1981) among residents of Fukuro and Tsukinoura districts (inhabited mainly by fishermen and their

families) in the Kumamoto Prefecture with age-matched residents of Minamata city (also in the

Kumamoto Prefecture) who died between 1972 and 1978. In this study, high exposure to methylmercury

was inferred from residence in a district believed to have higher intake of local seafood. By contrast, in

the 1984 study described above, high methylmercury exposure was inferred from a diagnosis of

Minamata disease. A total of 416 deaths were recorded in the Fukuro and Tsukinoura districts in

1970–1981, and 2,325 deaths were recorded in Minamata City in 1972–1978. No statistically significant

increase in the overall cancer mortality rate was observed; however, an increase in the mortality rate due

to liver cancer was observed (SMR, 207.3; 95% C.I. 116.0–341.9). Analysis of mortality by sex showed

a statistically significant increase in the rate of liver cancer only among males (SMR, 250.5; 95% C.I.

133.4–428.4). Males also had statistically significantly higher mortality due to chronic liver disease and

cirrhosis. The authors note that these results should be interpreted with caution because the population

of Fukuro and Tsukinoura districts had higher alcohol consumption and a higher prevalence of hepatitis

B (a predisposing factor for hepatocellular cancer). Interpretation of these results is also limited by an

incomplete description of the methodology used to calculate the SMRs; it is unclear whether the study

authors used appropriate methods to compare mortality data collected over disparate time frames (i.e. 12

years for exposed and seven years for controls).

In a study from Poland, Janicki et al. (1987) reported a statistically significant (p<0.02) increase

in the mercury content of hair in leukemia patients (0.92 ± 1.44 ppm; n=47) relative to that in healthy

unrelated patients (0.49 ± 0.41 μg/g; n=79). Similarly, the mercury content in the hair of a subgroup of

leukemia patients (0.69 ±0.75 μg/g; n=19) was significantly (p<0.05) greater than that in healthy

relatives who had shared the same residence for at least 3 years preceding the onset of the disease (0.43 ±

0.24 μg/g; n=52). When patients with specific types of leukemia were compared with the healthy

unrelated subjects (0.49 ± 0.41 μg/g; n=79), only those with acute leukemia (type not specified; 1.24 ±

1.93 μg/g; n=23) had a significantly increased hair mercury content. No significant differences in hair

mercury content were observed in 9 patients with chronic granulocytic leukemia or 15 patients with

chronic lymphocytic leukemia when compared to the unrelated, healthy controls. This study is of limited

use for cancer risk assessment because of the following: small size of population studied; inadequate

description of the leukemia patients or healthy controls (e.g., age distribution, length of residence in the
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region, criteria for inclusion in the study); uncertainty regarding the source of mercury exposure (the

authors presumed that exposure was the result of use of methylmercury-containing fungicides);

uncertainty regarding the correlation between the chronology of incorporation of mercury in the hair and

onset of the disease; and the failure to address exposure to other chemicals or adjust for other leukemia

risk factors. Furthermore, the variability of hair mercury content was large, and the mean hair mercury

levels were within normal limits for all groups. One cannot rule out the likelihood that the observed

correlation of leukemia incidence with mercury in hair is due to chance alone.

The carcinogenic effects of organomercury seed dressing exposure were investigated in a series

of case-control studies for incidence of soft-tissue sarcomas (Eriksson et al. 1981; Hardell and Eriksson,

1988; Eriksson et al. 1990) or malignant lymphomas (Hardell et al. 1981). These studies were conducted

in Swedish populations exposed to phenoxyacetic acid herbicides or chlorophenols (the exposures of

primary interest in the studies), organomercury seed dressings, or other pesticides. Exposure frequencies

were derived from questionnaires and/or interviews. Control groups from the same region of the country

were matched to cases based on vital status. There were 402 total cases of soft-tissue sarcoma, and

(among persons not exposed to phenoxyacetic acid herbicides) there were 128 cases of malignant

lymphoma. In each study, the odds ratio for exposure to organomercury in seed dressings and sarcoma or

lymphoma was either less than 1.0, or the range of the 95% confidence interval for the odds ratio

included 1.0; therefore, no association was indicated for organomercury exposure and soft-tissue sarcoma

or malignant lymphoma. The conclusions from these studies are limited, however, due to the study

subjects' likely exposures to the other pesticides and chemicals.

Table 3-52

Carcinogenic Effects of Methylmercury in Humans: Epidemiological Studies

Species/ Exposure Dose

No. per Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/334 NS NS No increase in cancer mortality among Minamata exposure Tamashiro et al. 1984

exposed (M+F), victims (i.e., with overt methylmercury poisoning). 

668 control Minamata disease was used as a surrogate for

methylmercury exposure.

Limitations: Exposure levels or number of undiagnosed

cases among controls not known.

Human/412 NS NS Increased incidence of liver cancer in males living in the Tamashiro et al. 1986

exposed (M+F) vicinity of Minamata Bay. 

Limitations: "Exposed" districts had higher alcohol

consumption and higher prevalence of hepatitis B.

Human/47 w/ NS NS Increased mercury in hair of leukemia patients; however, Janicki et al. 1987

leukemia (sex mean hair mercury levels in the leukemia patients was

not specified) within the normal range.

control 79 Limitations: Small study population; source of

methylmercury exposure not clear; failure to address other

leukemia risk factors or exposure to other chemicals.
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3.3.2.2 Animal Data

The results from three dietary studies in two strains of mice indicate that methylmercury is

carcinogenic. A fourth dietary study in mice, three dietary studies in rats and a dietary study in cats

failed to show carcinogenicity of methylmercury. Interpretation of two of the positive studies was

complicated by observation of tumors only at doses that exceeded the MTD. Interpretation of four nonpositive studies was limited because of deficiencies in study design or failure to achieve an MTD.

Methylmercuric chloride was administered in the diet at levels of 0, 0.4, 2, or 10 ppm (0, 0.03,

0.14 and 0.69 mg Hg/kg-day in males and 0, 0.03, 0.13, and 0.60 mg Hg/kg-day in females) to B6C3F1

mice (60/sex/group) for 104 weeks (Mitsumori et al. 1990). In high-dose males, a marked increase in

mortality was observed after 60 weeks (data were presented graphically; statistical analyses not

performed). Survival at study termination was approximately 50%, 60%, 60%, and 20% in control, low-,

mid-, and high-dose males, respectively, and 58%, 68%, 60%, and 60% in control, low-, mid-, and highdose females, respectively. The cause of the high mortality was not reported. At study termination, the

mean body weight in high-dose males was approximately 67% of controls and in high-dose females was

approximately 90% of controls (data presented graphically; statistical analyses not performed). Focal

hyperplasia of the renal tubules was significantly (p<0.01) increased in high-dose males (14/60; the

incidence was 0/60 in all other groups). The incidence of renal epithelial carcinomas (classified as solid

or cystic papillary type) was significantly (p<0.01) increased in high-dose males (13/60; the incidence

was 0/60 in all other groups). The incidence of renal adenomas (classified as solid or tubular type) was

also significantly (p<0.05) increased in high-dose males; the incidence was 0/60, 0/60, 1/60, and 5/60 in

control, low-, mid-, and high-dose males, respectively, and 0/60, 0/60, 0/60, and 1/60 in control, low-, 

mid-, and high-dose females, respectively. No metastases were seen in the animals. The incidences of a

variety of nonneoplastic lesions were increased in the high-dose rats including these: sensory

neuropathy, neuronal necrosis in the cerebrum, neuronal degeneration in the cerebellum, and chronic

nephropathy of the kidney. Males exhibited tubular atrophy of the testis (1/60, 5/60, 2/60, and 54/60 in

control, low-, 

mid-, and high-dose, respectively) and ulceration of the glandular stomach (1/60, 1/60, 0/60, and 7/60 in

control, low-, mid-, and high-dose males, respectively). An MTD was achieved in mid-dose males and

high-dose females. High mortality in high-dose males indicated that the MTD was exceeded in this

group.

Mitsumori et al. (1981) administered 0, 15, or 30 ppm of methylmercuric chloride (99.3% pure)

in the diet (0, 1.6 and 3.1 mg Hg/kg-day) to ICR mice (60/sex/group) for 78 weeks. Interim sacrifices of

up to 6/sex/group were conducted at weeks 26 and 52. Kidneys were microscopically examined from all

animals that died or became moribund after week 53 or were killed at study termination. Lungs from

mice with renal masses and renal lymph nodes showing gross abnormalities were also examined. 

Survival was decreased in a dose-related manner; at week 78 survival was 24/60, 6/60 and 0/60 in

control, low- and high-dose males, respectively, and 33/60, 18/60 and 0/60, in control, low- and highdose females, respectively (statistical analyses not performed). The majority of high-dose mice (51/60

males and 59/60 females) died by week 26 of the study. Examination of the kidneys of mice that died or

were sacrificed after 53 weeks showed a significant (p<0.001) increase in renal tumors in low-dose males

(13/16 versus 1/37 in controls). The incidence of renal epithelial adenocarcinomas in control and lowdose males was 0/37 and 11/16, respectively (p<0.001). The incidence of renal epithelial adenomas in

control and low-dose males was 1/37 and 5/16, respectively (p<0.01). No renal tumors were observed in

females in any group. No metastases to the lung or renal lymph nodes were observed. Evidence of

neurotoxicity and renal pathology was observed in the treated mice at both dose levels. The high

mortality in both groups of treated males and in high-dose females indicated that the MTD was exceeded

in these groups. 
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A follow-up study to the Mitsumori et al. (1981) study was reported by Hirano et al. (1986). 

Methylmercuric chloride was administered in the diet to ICR mice (60/sex/group) at levels of 0, 0.4, 2, or

10 ppm (0, 0.03, 0.15, and 0.73 mg Hg/kg-day in males and 0, 0.02, 0.11, and 0.6 mg Hg/kg-day in

females) for 104 weeks. Interim sacrifices (6/sex/group) were conducted at 26, 52, and 78 weeks. 

Complete histopathological examinations were performed on all animals found dead, killed in extremis,

or killed by design. Mortality, group mean body weights and food consumption were comparable to

controls. The first renal tumor was observed at 58 weeks in a high-dose male, and the incidence of renal

epithelial tumors (adenomas or adenocarcinomas) was significantly increased in high-dose males (1/32,

0/25, 0/29, and 13/26 in the control, low-, mid-, and high-dose groups, respectively). Ten of the 13

tumors in high-dose males were adenocarcinomas. These tumors were described as solid type or cystic

papillary types of adenocarcinomas. No invading proliferation into the surrounding tissues was seen. 

The incidence of renal epithelial adenomas was not significantly increased in males, and no renal

adenomas or adenocarcinomas were observed in any females. Focal hyperplasia of the tubular

epithelium was reported to be increased in high-dose males (13/59; other incidences not reported). 

Increases in nonneoplastic lesions in high-dose animals provided evidence that an MTD was exceeded. 

Nonneoplastic lesions reported as increased in treated males included the following: epithelial

degeneration of the renal proximal tubules; cystic kidney; urinary cast and pelvic dilatation; and

decreased spermatogenesis. Epithelial degeneration of the renal proximal tubules and degeneration or

fibrosis of the sciatic nerve were reported in high-dose females.

Table 3-53

Carcinogenic Effects of Methylmercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/strain NS/ 25 2 yr 0, 0.004, 0.02, Tumors at comparable incidence in all groups Verschuuren et al.

M, 25 F ad lib in feed 0.1 Limitations: Small sample size; failure to achieve MTD 1976

BML avg: 850 μg/L in blood at 0.004, 6,500 μg/L at

0.02, and 36,000-39,000 μg/L at 0.1

Rat/ 130 wk 0.011, 0.05, No increase in tumor incidence Mitsumori et al.

Sprague ad lib in FEFD 0.28 (M); 0.014, 1983, 1984

Dawley/56 M, 56 0.064, 0.34 (F)

F

Mice/Swiss/ weaning until 0, 0.19, 0.19- No increase in gross tumor incidence Schroeder and

54 M, 54 F death in 0.95 Limitation: Histological examination not performed. Mitchener 1975

drinking water (MMA)

Mouse/ICR/ 78 wk ad lib 0, 1.6, 3.1 Increased incidence renal adenomas and adenocarcinomas Mitsumori et al. 1981

60 M, 60 F in feed in low-dose males.

Limitations: Very poor survival in both male dose

groups.

Mouse/ICR/ 104 wk ad lib 0, 0.02, 0.03, Incidence of renal epithelial adenocarcinoma significantly Hirano et al. 1986

60 M, 60 F in feed 0.11, 0.15, increased in males at 0.73; not invasive.

0.6, 0.73 Limitations: MTD exceeded (including severe renal

damage in high-dose males)

Mouse/ 2 yr 0.03, 0.14, 0.69 Renal epithelial carcinomas and adenomas in males at Mitsumori et al. 1990

B6C3F1/ ad lib in feed (M); 0.03, 0.13, 0.69.

60 M, 60 F 0.6 (F) Limitation: MTD exceeded in high-dose males.
Table 3-53 (continued)

Carcinogenic Effects of Methylmercury in Animals: Oral Exposure

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

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Mice/Swiss/ 15 wk ad lib 0, 0.03, 0.07, Number of lung adenomas/mouse and tumor size/mouse Blakley 1984

NS in drinking 0.27 increased with dose

water (MMC)

Cat/domestic/ 2 yr 0, 0.0084, 0.02, No increase in tumor incidence Charbonneau et al.

4-5 M, 4-5 F ad lib in feed 0.046, 0.074, Limitations: Small group size, short exposure duration, 1976

0.176 no pathological data for 3 lowest doses.

No increase in tumor incidence was observed in a study using white Swiss mice (Schroeder and

Mitchener 1975). Groups of mice (54/sex/group) were exposed from weaning until death to

methylmercuric acetate in the drinking water at two doses. The low-dose group received 1 ppm

methylmercuric acetate (0.19 mg Hg/kg-day). The high-dose group received 5 ppm methylmercuric

acetate (0.95 mg Hg/kg-day) for the first 70 days and then 1 ppm, thereafter, due to high mortality (21/54

males and 23/54 females died prior to the dose reduction). Survival among the remaining mice was not

significantly different from controls. Significant (p<0.001) reductions in body weight were reported in

high-dose males (9–15% lower than controls) and high-dose females (15–22% lower than controls)

between 2 and 6 months of age. Mice were weighed, dissected, gross tumors were detected, and some

sections were made of heart, lung, liver, kidney and spleen for microscopic examination. No increase in

tumor incidence was observed. This study is limited because complete histological examinations were

not performed, and pathology data other than tumor incidence were not reported.

Mitsumori et al. (1983, 1984) conducted a study in Sprague-Dawley rats. They administered

diets containing 0, 0.4, 2, or 10 ppm of methylmercuric chloride (0, 0.011, 0.05, and 0.28 mg Hg/kg-day

in males; 0, 0.014, 0.064 and 0.34 mg Hg/kg-day in females) to Sprague-Dawley rats (56

animals/sex/group) for up to 130 weeks. Interim sacrifices of 10/group (either sex) were conducted at

weeks 13 and 26 and of 6/group (either sex) at weeks 52 and 78. Mortality was increased in high-dose

males and females. At week 104, survival was approximately 55%, 45%, 75% and 10% in control, low-,

mid-, and high-dose males, respectively, and 70%, 75%, 75% and 30% in control, low-, mid-, and highdose females, respectively (data presented graphically). Body weight gain was decreased in high-dose

animals (approximately 20–30%; data presented graphically). No increase in tumor incidence was

observed in either males or females. Noncarcinogenic lesions that were significantly increased (p< 0.05)

in high-dose rats included the following: degeneration in peripheral nerves and the spinal cord (both

sexes); degeneration of the proximal tubular epithelium of the kidney (both sexes); severe chronic

nephropathy (females); parathyroid hyperplasia (both sexes); polyarteritis nodosa and calcification of the

abdominal arterial wall (females); bone fibrosis (females); bile duct hyperplasia (males); and

hemosiderosis and extramedullary hematopoiesis in the spleen (males). In addition, mid-dose males

exhibited significantly increased degeneration of the kidney proximal tubular epithelium and hyperplasia

of the parathyroid. An MTD was achieved in mid-dose males and in high-dose females; the MTD was

exceeded in high-dose males.

No increase in tumor incidence or decrease in tumor latency was observed in another study using

rats (strain not specified) (Verschuuren et al. 1976). Groups of 25 female and 25 male rats were
3-72

administered methylmercuric chloride at dietary levels of 0, 0.1, 0.5 and 2.5 ppm (0, 0.004, 0.02 and 0.1

mg Hg/kg-day) for 2 years. No significant effects were observed on growth or food intake except for a

6% decrease (statistically significant) in body weight gain at 60 weeks in high-dose females. Survival

was 72%, 68%, 48% and 48% in control, low-, mid- and high-dose males, respectively; and 76%, 60%,

64% and 56% in control, low-, mid- and high-dose females, respectively (statistical significance not

reported). Increases in relative kidney weights were observed in both males and females at the highest

dose. No effects on the nature or incidence of pathological lesions were observed, and tumors were

reported to have been observed with comparable incidence and latency among all of the groups. This

study was limited by the small sample size and failure to achieve an MTD.

No tumor data were reported in a study using Wistar rats (Munro, 1980). Groups of 50 Wistar

rats/sex/dose were fed diets containing methylmercury; doses of 2, 10, 50, and 250 micrograms Hg/kgday were fed for 26 months. High-dose female rats exhibited reduced body weight gains and showed

minimal clinical signs of neurotoxicity; however, high-dose male rats showed overt clinical signs of

neurotoxicity, decreased hemoglobin and hematocrit values, reduced weight gains and significantly

increased mortality. Histopathologic examination of the high-dose rats of both sexes revealed

demyelination of dorsal nerve roots and peripheral nerves. Males showed severe dose-related kidney

damage, and females had minimal renal damage.

No increase in tumor incidence was observed in a multiple generation reproduction study using

Sprague-Dawley rats (Newberne et al. 1972). Groups of rats (30/sex) were given semisynthetic diets

supplemented with either casein or a fish protein concentrate to yield dietary levels of 0.2 ppm

methylmercury (0.008 mg Hg/kg-day). Another group of controls received untreated rat chow. Rats that

received diets containing methylmercury during the 2-year study had body weights and hematology

comparable to controls. Detailed histopathologic analyses revealed no lesions of the brain, liver, or

kidney that were attributable to the methylmercury exposure. Mortality data were not presented. 

Interpretation of these data is limited by the somewhat small group sizes and failure to achieve an MTD.

No increase in tumor incidence was observed in a study using random-bred domestic cats

(Charbonneau et al. 1976). Groups of cats (4–5/sex/group) were given doses of 0.0084, 0.020, 0.046,

0.074 or 0.176 mg Hg/kg-day either as methylmercury-contaminated seafood or as methylmercuric

chloride in the diet for up to two years. Controls were estimated to have received 0.003 mg Hg/kg-day. 

Food consumption and body weight were not affected by treatment with methylmercury. Due to

advanced signs of neurotoxicity (loss of balance, ataxia, impaired gait, impaired reflexes, weakness,

impaired sensory function, mood change and tremor), cats at the highest dose tested were sacrificed after

approximately 16 weeks, and cats at the next highest dose were sacrificed after approximately 54–57

weeks. Cats at the next highest dose generally exhibited mild neurological impairment (altered hopping

reaction and hypalgesia). One cat at this dose was sacrificed after 38 weeks because of neurotoxicity,

and one cat died of acute renal failure after 68 weeks. Cats at the two highest doses had pathological

changes in the brain and spinal cord, but no histopathological changes were noted in other tissues

examined. Interpretation of the results of this study is limited because of the small group sizes, early

sacrifice of cats at the two highest dose levels and no available data regarding pathological changes in

cats at the three lowest dose levels. This study was also limited by its short duration when compared to

the lifespan of a cat.

Blakley (1984) administered methylmercuric chloride to female Swiss mice (number/group not

specified) in drinking water at concentrations of 0, 0.2, 0.5 or 2.0 mg/L for 15 weeks. This corresponded

to approximately 0, 0.03, 0.07 and 0.27 mg Hg/kg-day. At the end of week 3, a single dose of 1.5 mg/kg

of urethane was administered intraperitoneally to 16–20 mice/group. No effects on weight gain or food

consumption were observed. Lung tumor incidence in mice not administered urethane (number/group
3-73

not specified) was less than 1 tumor/mouse in all groups. Statistically significant trends for increases in

the number and size of lung adenomas/mouse with increasing methylmercury dose were observed; the

tumor number/mouse was 21.5, 19.4, 19.4 and 33.1 in control, low-, mid- and high-dose mice,

respectively, and the tumor size/mouse was 0.70, 0.73, 0.76 and 0.76 mm in control, low-, mid- and highdose mice, respectively. The study authors suggest that the increase in tumor number and size may have

been related to immunosuppressive activity of methylmercury. It should be noted that this is considered

a short term assay and that only pulmonary adenomas were evaluated.

3.3.3 Other Data

3.3.3.1 Death

Methylmercury is a potent toxin that causes impairment of the CNS and developmental toxicity

in humans. Ingestion of methylmercury from treated seed grain or contaminated fish has resulted in

death. An outbreak of methylmercury poisoning in Iraq caused deaths in people who consumed

methylmercury from bread made with grain treated with a fungicide (Al-Saleem and the Clinical

Committee on Mercury Poisoning 1976; Bakir et al. 1973). The deaths were attributed to impaired CNS

function. A syndrome known as Minamata disease has been characterized by nervous system impairment

following consumption of methylmercury-contaminated fish from Minamata Bay in Japan. Symptoms of

Minamata disease include the following: prickling; tingling sensation of extremities; impaired peripheral

vision, hearing, taste and smell; slurred speech; muscle weakness; irritability; memory loss; depression;

and sleeping difficulties (Kutsuna 1968; Takeuchi et al. 1962; Tsubaki and Takahashi 1986). Deaths

from Minamata disease can be broken into two categories: deaths occurring from the beginning of the

outbreak (1954) to 1969, and deaths occurring from 1970 to 1980 (Tamashiro et al. 1984). Over half of

the deaths in the first group were attributed to Minamata disease and/or noninflammatory disease of the

central nervous system, or pneumonia, whereas deaths in the second group were attributed to

cerebrovascular disease with underlying Minamata disease. The mean age at death for the first group

was 45.4 years for males and 26.4 years for females, and the mean age at death for the second group was

70.0 years for males and 72.7 years for females (Tamashiro et al. 1984).

Table 3-54

Lethality of Methylmercury in Humans: Case Study of Oral Exposure

Species/ Exposure Dose

No. per Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/6,530 43-68 d 0.71-5.7 (est.) Of 6,350 cases admitted to hospitals, 459 died after eating Bakir et al. 1973

both sexes (feed) bread made from grain treated with methylmercury

fungicide

BML: <100-5,000 μg/L in blood

Human/1,422 NS NS Of 1,422 patients from the Minamata disease outbreak in Tamashiro et al.

both sexes 1959, 378 died by 1980. 1984

Limitation: exposure data limited

BML not reported

Very little information regarding death after inhalation exposure to methylmercury was located. 

One study reported a man who died after being exposed for three years to alkylmercury particles from

seed dressings (Hook et al. 1954). Prior to death, the man experienced increased symptoms of

neurotoxicity. A case study reported on the deaths of two women exposed to diethylmercury vapors for
3-74

3–5 months (Hill 1943). Gastrointestinal effects and neurological symptoms occurred prior to deaths. 

No other human studies were located regarding death after inhalation exposure to methylmercury.

Table 3-55

Lethality of Methylmercury in Humans: Case Studies of Inhalation Exposure

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/2 F 3-5 mo NS Death following exposure to diethylmercury vapors Hill 1943

(occup) Limitation: Case study; concomitant dermal exposure likely;

limited exposure data

BML not reported

Human/1 M 3 yr NS Death following exposure to pesticide containing Hook et al. 1954

(occup) methylmercury

Limitations: Case study; concomitant dermal exposure likely;

limited exposure data

Range: 500-640 μg/L in urine

Mice given a single oral dose of methylmercury had an increased incidence of death compared to

controls (Yasutake et al. 1991). Male mice appear to be more sensitive to the effects of methylmercury

than females, possibly due to the effect of mercury on the male kidneys. Mice exposed for 26 weeks to

3.1 mg Hg/kg-day as methylmercury in drinking water also showed an increase in mortality compared to

controls (51/60 males and 59/60 females of exposed group died versus 1/60 males and 1/60 females in

controls) (Mitsumori et al. 1981). Longer studies (78 and 104 weeks) confirm that methylmercury causes

significantly increased mortality in mice compared to controls (Mitsumori et al. 1981, 1990). No animal

studies were located on death after inhalation exposure to methylmercury.

Table 3-56

Lethality of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Mouse/ICR/60 78 wk 0, 1.6, 3.1 51/60 males and 59/60 females receiving 3.1 mg/kg/d Mitsumori et al.

M, 60 F ad lib in feed (MMC) died by week 26, vs 7 males and 6 females at 1.6, and 1 1981

control males and 1 control female; death at 52 wk was

also elevated at 1.6

Limitation: No statistical analysis

BML not reported

Mouse/ 104 wk 0, 0.03, 0.13, 50/60 males treated with 0.69 mg/kg/d died vs. 31/60 Mitsumori et al.

B6C3F /60 M, ad lib in feed 0.14, 0.60, 0.69 control males; survival of females and males at lower 1990 1

60 F (MMC) doses was unaffected

BML not reported
Table 3-56 (continued)

Lethality of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

3-75

Mouse/ Once 4, 8, 16, 24, 32, 4/6 males died (LOAEL = 16); LOAEL for females was Yasutake et al. 1991

C57BL/6 M, 6 40 40 (4/6 died); no statistical analysis or LD calculated

F Limitation: small number of animals tested50

BML: 2.45 μg/g in kidney of males at 16 mg/kg

3.3.3.2 Neurological

The nervous system is the primary target organ for methylmercury toxicity. Information from the

large-scale poisonings in Japan (Niigata and Minamata) and Iraq provide substantial information

regarding the neurotoxicity of methylmercury in humans (Bakir et al. 1973, 1980; Berglund et al. 1971;

Harada 1978; Marsh et al. 1987; Rustam and Hamdi 1974). In Japan, poisonings occurred between 1953

and 1960 when people consumed seafood that had been contaminated by methylmercury released by a

chemical plant into Minamata Bay and the Agano river near Niigata. In Iraq, poisonings occurred in the

winter of 1971 to 1972 when people ate bread made from seed grain that had been treated with a

mercury-containing fungicide. In all of these episodes, neurotoxicity was the most prominent effect

observed in the exposed populations. In the Iraqi incident, more than 6000 patients were hospitalized,

and more than 500 deaths occurred, usually due to CNS failure. 

The least severely affected persons from the poisonings in Japan and Iraq experienced numbness

or tingling (paresthesia) of the extremities and/or perioral area. Additional symptoms frequently

experienced by more severely affected individuals included the following: ataxia (gait impairment

ranging from mild incoordination or unsteadiness to complete inability to walk); blurred vision;

constriction of visual fields (in extreme cases blindness); slurred speech; and hearing difficulties

(deafness in extreme cases). Less frequently observed symptoms associated with the methylmercury

poisonings included tremors, muscular weakness, abnormal reflexes, increased muscle tone, and clouded

memory or stuperousness. A long latent period (16–38 days in the Iraqi episode and up to several years

in the Japanese episodes) between exposure and onset of symptoms of neurotoxicity was observed. The

cause for the latent period is unknown. It is thought that latency may be related to cellular repair

mechanisms dealing with damage from lipid peroxidation. At the point when repair processes are

overwhelmed tissue damage and accompanying symptoms become apparent. The possible ameliorating

effect of selenium in the diet has also been hypothesized to play a part in latency.

Similar neurological symptoms have been observed in persons ingesting meat contaminated with

ethylmercuric chloride (Cinca et al. 1979). Two boys who ultimately died from exposure exhibited

neurological signs including gait disturbance, ataxia, dysarthria, dysphagia, aphonia, hyperactive tendon

reflexes, hypotonia, mydriasis and agitation. In the surviving members of the family, ataxia, gait

impairment, spasticity, drowsiness, intention tremor, agitation, speech difficulties and visual disturbances

were reported. All effects except the narrowing of the visual fields disappeared after exposure

termination. 
3-76

Histopathologic analyses of nervous system tissue taken from poisoning victims show neuronal

degeneration in the cerebrum and cerebellum (Bakir et al. 1980; Swedish Expert Group 1971; Takeuchi

et al. 1962). In the cerebral cortex, the calcarine area was most regularly affected with varying degrees

of damage in the pre- and postcentral cortices, superior temporal gyrus, and basal ganglia. In the

cerebellar cortex, granule cell loss predominated, but this was usually less severe than cerebral damage. 

An autopsy of two boys who ingested ethyl mercury contaminated meat revealed nerve cell loss and glial

proliferation in the cerebral cortex, demyelination, granule cell loss in the cerebellum, and motor neuron

loss in the ventral horns of the spinal cord (Cinca et al. 1979). Less information is available regarding

the histopathology of peripheral nerve involvement, but sural nerves taken from two victims of the

Minamata episode showed evidence of peripheral nerve degeneration and regeneration (Miyakawa et al.

1976). Fourteen Iraqi patients who developed ataxia and "pins and needles" and could not perform heelto-toe walk were examined for impaired peripheral nerve function (Von Burg and Rustam 1974a, 1974b). 

Determinations of motor and sensory conduction velocities, sensory threshold and latency, reflex of the

tibial nerve and myoneural transmission were performed, but there were no statistical significances

between exposed and unexposed control groups; the mean values of the experimental group, however,

were somewhat lower than those of the controls. There was also no consistent correlation between

clinical or electrophysiological observation on the peripheral nervous system and blood mercury levels. 

In two patients who were hospitalized 10 days after ingestion of ethyl mercury-contaminated meat,

sensory nerve conduction velocity was decreased immediately after admission but was found to be

normal six months later (Cinca et al. 1979).
3-77

Table 3-57

Neurotoxicity of Methylmercury in Humans: Case Studies of Oral Exposure

Species/ Exposure Dose

No. per Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/14 NS NS Ataxia; impaired heel-to-toe walk; complaints of "pins and Von Burg and

cases needles". Sensory and motor peripheral nerves were not Rustam 1974a,

affected. Clinical and electrophysiological observations 1974b

did not correlate with blood concentration

Limitation: Exposure concentration and duration not

known

BML: blood Hg levels were 138-878 μg/L

Human/No. NS NS NS Paresthesia/numbness; constriction of visual field; Harada 1978

incoordination; difficulty speaking; tremor in consumers

of contaminated fish

Limitation: Limited details reported

BML not reported

Human/2 M, 2 Once NS Gait disturbance, ataxia, dysarthria, speech difficulties, Cinca et al. 1979

F (ethyl mercury visual disturbances, hyperactive tendon reflexes,

chloride) mydriasis, agitation, coma; nerve degeneration in cerebral

cortex, cerebellum, and ventral horns of spinal cord;

decreased sensory nerve conduction velocity. Ingestion of

ethyl mercury chloride-contaminated meat.

Limitation: Exposure concentration not known

BML: hair Hg levels of 152-542 μg/g

Human/6530 43-68 d 0.71-5.7 (est.) Paresthesia/numbness in extremities and perioral area; Bakir et al. 1973,

cases both ataxia; constriction of visual field or blindness; slurred 1980

sexes speech; hearing difficulties following ingestion of grain

contaminated with methylmercury. Incidence and severity

of effect correlated with blood concentration

BML: Total body burden 50 mg at time of onset

Human/81 F <5 mo NS Paresthesia and "other neurological symptoms" Marsh et al. 1987

(MMC) BML Range: 1-674 μg/g Hg in hair; one woman with 14

μg/g (maximum in strand) had paresthesia and a woman

with 10 μg/g had other symptoms. However, others with

levels as high as 600 μg/g had no symptoms. (This is a

follow-up study to Bakir et al. 1980)

There are two case studies that report neurotoxicity in humans following inhalation of

methylmercury (Hook et al. 1954; Hunter et al. 1940); however, no quantitative data were available. The

two studies described in Table 3-58 demonstrated the spectrum of neurotoxic effects that occur following

occupational exposure to methylmercury. Weiss and Simon (1975) have suggested that such changes in

function in the general population, particularly at relatively low doses, may not be clinically detectable as

a loss of function but may be unmasked by the normal processes of aging.
3-78

Table 3-58

Neurotoxicity of Methylmercury in Humans: Case Studies of Inhalation Exposure

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitation/BML Reference 3

Human/5 M 5 mo-2 yr NS Tingling of limbs; unsteady gait; difficulty performing fine Hunter et al. 1940

(occup) movements; constricted visual field following exposure to

methylmercury nitrate, methylmercury iodide, of

methylmercury phosphate in chemical factories

Limitations: Case studies; concomitant dermal exposure and

exposure to other chemicals likely; limited exposure data

BML not reported

Human/1 M 3 yr NS Weakness in arms and legs; irregular EEG; sensory and speech Hook et al. 1954

(occup) disorders following exposure to pesticide containing

methylmercury

Limitations: Case study; concomitant dermal exposure likely;

limited exposure data

BML Range: 500-640 μg/L in urine

As a result of the methylmercury poisonings in Japan and Iraq, substantial information on the

neurotoxicity of methylmercury has been generated from animal studies. Relatively brief, high level

exposures in rats have been shown to cause characteristic signs of neurotoxicity (flailing and hindlimb

crossing when the animal is lifted by the tail) and neuronal degeneration in the cerebellum, cerebral

cortex and dorsal root ganglia (Inouye and Murakami 1975; Leyshon and Morgan 1991; Magos et al.

1985; Yip and Chang 1981). As with humans there is a latency period; the effects frequently are not

observed or do not show maximal severity until several days after the cessation of dosing. In an acute

study, exposure of rats to a single gavage dose of 19.9 mg Hg/kg as methylmercuric chloride resulted in

impaired open-field tests such as decreases in standing upright, area traversed and activity compared to

the control group (Post et al. 1973). Animals were lethargic and ataxic initially, but symptoms

disappeared within 3 hours.

Longer-term, lower-level exposures revealed that evidence of neuronal degeneration may be

observed prior to the onset of overt signs of toxicity. Degeneration in the cerebellum was found in rats

given 10 mg Hg/kg as methylmercuric chloride once every 3 days for 15 days (Leyshon and Morgan

1991) while severe degenerative changes in the dorsal root fibers were observed in rats given 1.6 mg

Hg/kg-day as methylmercuric chloride for 8 weeks (Yip and Chang 1981). Munro et al. (1980) observed

demyelination of dorsal nerve roots and damage in sciatic nerves with oral exposure to 0.25 mg Hg/kgday as methylmercuric chloride for up to 26 months. In mice given 1.9 mg Hg/kg-day as methylmercury,

cerebellar lesions were observed as early as eight days after the start of dosing, but changes in motor

activity did not develop until 24 weeks of exposure (MacDonald and Harbison 1977). Similarly, cats

receiving methylmercury in the diet for 11 months displayed degenerative changes in the cerebellum and

cerebral cortex, but incoordination or weakness was observed in only a small number of the animals with

histopathological changes (Chang et al. 1974).

The molecular basis for methylmercury neurotoxicity is likely to be complex and multifactorial. 

The broad affinity of mercury for -SH groups leads to membrane, enzyme and cytoplasmic organelle

interaction. Major mechanistic pathways have been proposed to include the following: inhibition of

macromolecular metabolism, especially that of protein translation and nucleic acid biogenesis; oxidative
3-79

injury; disturbance in Ca hemostasis; aberrant protein phosphorylation. The mechanisms underlying 2+

inhibition of protein and RNA synthesis are multiple. Depending upon the systems used with in vitro, in

vivo or neuronal cell suspensions, evidence for inhibition of translation associated with a change in

ATP/ADP concentration has been found. On the other hand, direct inhibition of elongation was

documented secondary to the selective inhibition of certain aminoacyl-tRNA synthetase (Cheung and

Verity, 1985). Syversen (1977) investigated the effects of methylmercury on protein synthesis in rats

using techniques which allowed analysis of different cell populations from the central nervous system. 

Results of this study indicated selective irreversible damage to granule cells of the cerebellum, whereas

damage to the other neurons, such as Purkinje cells was reversible. Such selectivity of toxicity is a

feature of the neuronal loss seen in human and experimental disease. Methylmercury has also been

suggested to cause neuronal degeneration by promoting the formation of reactive oxygen species (Ali et

al. 1992; Le Bel et al. 1990, 1992; Verity and Sarafian 1991). While contributory, such oxidative injury

does not appear primary to the site of toxicity as appropriate protective measures blocking oxidative

stress and lipoperoxide formation are only minimally cytoprotective.

A recent review by Atchison and Hard (1994) discusses several proposed mechanisms of action

of methylmercury on Ca hemostasis and ion channel function. Individual studies have demonstrated 2+

that the neuromuscular actions of methylmercury occur predominantly at the presynaptic site (Atchison

et al. 1984). Methylmercury may interfere with acetylcholine neurotransmitter release and subsequently

synaptic transmission (Atchison et al. 1986; Barrett et al. 1974; Schafer et al. 1990; Schafer and Atchison

1989, 1991). Finally, Sarafian (1993) demonstrated that the methylmercury-induced stimulation of

protein phosphorylation in cerebellar granule cell culture is coupled to Ca uptake, changed intracellular 2+

Ca hemostasis and inositol phosphate metabolism. These latter observations invoke the activation of 2+

the protein kinase C pathway.

Cats and monkeys appear to be more sensitive to the neurotoxic effects of methylmercury than

rodents. Long-term studies in primates and in cats have shown neurological impairment at doses as low

as 0.05 mg Hg/kg-day. In cats, mild impairment of motor activity and decreased pain sensitivity was

observed at 0.046 mg Hg/kg-day as methylmercury after 60 weeks of exposure (Charbonneau et al.

1976). In cynomolgus monkeys given methylmercury from birth until approximately 7 years of age,

impairment of spatial visual function was observed after 3 years, and decreased fine motor performance,

touch and pinprick sensitivity and impaired high frequency hearing were observed six to seven years after

cessation of dosing (Rice 1989b; Rice and Gilbert 1982, 1992). Exposure of cynomolgus monkeys to

0.03 mg Hg/kg-day as methylmercury for approximately 4 months caused no detectable changes in motor

activity or effects on vision or hearing, but degenerative changes were observed in neurons of the

calcarine cortex and sural nerve when these were examined electron microscopically (Sato and Ikuta

1975). At higher doses (0.08 mg Hg/kg-day), slight tremor, motor incoordination and blindness were

observed in Macaca fascicularis monkeys after four months of exposure (Burbacher et al. 1988).

The developing organism is generally at higher risk of neurotoxicity than adults. The section on

developmental effects of methylmercury lists studies wherein animals were observed with neurological

or neurobehavioral deficits as a consequence of in utero or perinatal methylmercury exposure.
3-80

Table 3-59

Neurotoxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/Wistar/ 0–12 or 2, 4 Hindlimb crossing (LOAEL = 4) after 0–12 days Inouye and

10 F 12–20 d, 1x/d (MMC) BML not reported Murakami 1975

(gavage)

*

Rat/Wistar/ up to 26 mo 0.002, 0.01, Ruffled fur, loss of balance, hindlimb crossing, paralysis Munro et al. 1980

50 F, 50 M ad lib in feed 0.05, 0.25 (LOAEL = 0.25) after 6 mo (males more affected);

(MMC) demyelination of dorsal nerve roots and damage in teased *

sciatic nerves at 0.25

Avg. BML at 0.25: 115 ppm in blood

Rat/Charles 8 wk 0, 1.6 Degeneration of dorsal root fiber Yip and Chang 1981

River/6 M 7 d/wk (MMC) BML not reported

1 x/d

(gavage)

*

Rat/Wistar/24 5 d 8 Cerebellar granule cell and dorsal root ganglion cell Magos et al. 1985

M, 18 F 1 x/d (MMC) degeneration; flailing and hind leg crossing following

(gavage) administration of methylmercuric chloride

*

Limitations: Only one level tested; no controls

Avg BML: 150,000 μg/L in blood

Rat/Wistar/15 5 x/15 d 0, 10 Granule cell degeneration in cerebellum Leyshon and

M (gavage) (MMC) BML: 60 μg/g dry cerebellar weight Morgan 1991 *

Swiss origin 28 wk 1.9, 9.5 Ataxia; degenerative changes of Purkinje cells; granule MacDonald and

Mouse (ad lib (MMC) cell loss in cerebellum; (LOAEL = 1.9) Harbison 1977

M drinking BML not reported

water)

*

Cat/Breed 11 mo 0, 0.015 Degeneration of cerebellum and cerebral cortex; necrosis Chang et al. 1974

NS/15-16 both (ad lib in (MM) of dorsal root ganglia of kittens fed mercury-contaminated

sexes feed) tuna

BML not reported

Cat/Breed 2 yr 0.003, 0.008, Impaired hopping reaction; decreased pain sensitivity; Charbonneau et al.

NS/8-10 NS 7 d/wk 0.020, 0.046, degeneration of dorsal root ganglia (LOAEL = 0.046) 1976

(feed) 0.074, 0.176 Avg BML: 9,000 μg/L in blood at 0.046 mg/kg-day

(MMC)*

Monkey/ 36–132 d 0.02, 0.03, Atrophy of neurons in calcarine cortex; focal degeneration Sato and Ikuta 1975

Macaca 1 x/d 0.04, 0.07, 0.21 in sural nerves (LOAEL=0.03); ataxic gait, myoclonic

fascicularis/1-2 (feed) seizures at 0.21 mg/kg-day

both sexes Limitation: small number of animals tested

BML: Maximal at 0.03 mg/kg-day of 460 μg/L in blood

and 62 μg/g in hair

Monkey/ 90-270 d 1 for 5 doses, Tremor; visual impairment (LOAEL = 0.5 mg/kg) Evans et al. 1977

Macaca 1 x/wk then Limitations: Small number of animals tested, limited

artoides, (gavage) 0.4, 0.5, 0.6 description of effects

Macaca Avg BML: 2,900 μg/L in blood

nemestrina/2

both sexes
Table 3-59 (continued)

Neurotoxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

3-81

Monkey/ 3-4 yr 0, 0.05 Spatial visual impairment Rice and Gilbert

Macaca 7 d/wk (MMC) Limitation: One dose level tested 1982

fascicularis/5 1 x/d BML: 600-900 μg/L in blood

exposed, 2 (NS)

control (sex

NS)

*

Monkey/ 3 yr 0, 0.04, 0.06, Slight tremor; motor incoordination; blindness (LOAEL = Burbacher et al.

Macaca 1 x/d 0.08 0.04) following administration of methylmercury 1988

fascicularis/7-8 (oral route (MMC) hydroxide; time to onset was 177-395 d

F NS) Avg BML: 2,030 μg/L in blood at highest dose

*

Monkey/ 6.5-7 yr 0, 0.05 Six years after end of dosing (follow-up study to Rice and Rice 1989b; Rice

Macaca 7 d/wk (MMC) Gilbert 1982): decreased fine motor performance; and Gilbert 1992

fascicularis/4 1 x/d diminished touch and pinprick sensitivity; impaired high

M, 1 F (capsule; frequency hearing (p<0.05)

exposed, 1 M, 2 gavage) Limitations: small number of animals tested; one dose

F controls level tested

*

BML: Not detectable at time of testing

MMC = methylmercuric chloride *

3.3.3.3 Renal

No studies were located regarding the renal toxicity of methylmercury in humans following oral

exposure. Renal histopathology and decreased function have been observed following acute or chronic

oral exposure of rats and mice to methylmercury. Renal tubule vacuolation was observed in rats

receiving 8 mg Hg/kg-day for 5 days (Magos et al. 1985), and decreased phenolsulfonphthalein excretion

occurred in male mice receiving a single dose of 16 mg Hg/kg-day or greater and females at 32 mg

Hg/kg-day or greater as methylmercuric chloride (Yasutake et al. 1991). Chronic nephropathy, including

epithelial degeneration of proximal tubules and interstitial fibrosis, was observed at longer durations

(Fowler 1972; Hirano et al. 1986; Mitsumori et al. 1990). Males were more sensitive than females to

renal effects (Mitsumori et al. 1990).
3-82

Table 3-60

Renal Toxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/Wistar/3 12 wk 0, 0.08 (M) Cytoplasmic mass in proximal tubule cells Fowler 1972

M, 6 F ad lib in feed 0, 0.09 (F) Limitation: Only one level tested; small number of treated

exposed/16 (MMC) animals

controls (sex BML not reported

NS)

Rat/Wistar/24 5 d 8 Renal tubule vacuolation and dilation Magos et al. 1985

M, 18 F 1 x/d Limitation: One level tested, no controls

(gavage) Avg. BML: 150,000 μg/L in blood

Mouse/ICR/60 26 wk 0, 0.03, 0.15, Toxic epithelial degeneration of renal proximal tubules Hirano et al. 1986

M, 60 F ad lib in feed 0.72 (M); 0.02, (LOAEL = 0.62 F; 0.72 M)

0.11, 0.62 (F) BML not reported

Mouse/ 104 wk 0, 0.03, 0.14, Chronic nephropathy (epithelial cell degeneration, Mitsumori et al. 

B6C3F /60 M, ad lib in feed 0.68 (M); 0.03, regeneration of proximal tubules, interstitial fibrosis) in 1990 1

60 F 0.13, 0.6 (F) males at 0.14 and in females at 0.60 (p<0.01)

(MMC) BML not reported

Mouse/ Once 4, 8, 16, 24, 32, Decreased phenolsulfonphthalein excretion and increased Yasutake et al. 1991

C57BL/6 M, 6 (gavage) 40 (MMC) serum creatinine in males (LOAEL = 16 in males, 32 in

F females); swollen epithelial cells in proximal tubules

Limitation: No statistical analysis; small number of

treated animals

BML: 2.45 μg/g in kidneys of males and 1.9 μg/g in

kidneys of females at 16 mg/kg

3.3.3.4 Cardiovascular

Only one study was located regarding the cardiovascular toxicity of methylmercury in humans. 

Hook et al. (1954) reported two men with elevated blood pressure after inhalation exposure to organic

mercury particulates from seed dressings. Other neurotoxic effects were also present at the time of

examination, and one man subsequently died.
3-83

Table 3-61

Cardiovascular Toxicity of Methylmercury in Humans: Case Study

Species/ Exposure Dose

No. per Sex Duration (mg/m ) Effects/Limitations/BML Reference 3

Human/1 M 3 yr NS Elevated blood pressure Hook et al. 1954

(occup) Limitations: Case study; concomitant dermal exposure likely

BML Range: 500-640 μg/L in urine

Very little information was located regarding the effects of oral methylmercury exposure on the

cardiovascular system. Rats given two daily doses of methylmercuric chloride exhibited decreases in

heart rates (Arito and Takahashi 1991). Rats treated with methylmercuric chloride for one month had

increased systolic blood pressures beginning 42 days after cessation of dosing (Wakita 1987). This effect

persisted for more than a year.

Table 3-62

Cardiovascular Toxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitation/BML Reference

Rat/Wistar/10 23-28 d 0.4, 1.2 Increased systolic pressure beginning 42 d after the end of Wakita 1987

(sex NS) 7 d/wk (MMC) treatment (p<0.05)

(gavage) BML not reported

Rat/Sprague- 2 d 12 Decreased heart rate (p<0.05) Arito and Takahashi

Dawley/5-6 1 x/d (MMC) Limitation: Only one dose tested for this parameter 1991

(sex NS) (gavage) BML: 10 μg/g in brain

3.3.3.5 Gastrointestinal

No information was located regarding the gastrointestinal toxicity of methylmercury in humans. 

Only one study was located regarding the gastrointestinal toxicity of methylmercury following oral

exposure in animals. Mitsumori et al. (1990) reported an increased incidence of stomach ulceration in

mice following a 2-year exposure to 0.69 mg Hg/kg-day as methylmercuric chloride in drinking water.
3-84

Table 3-63

Gastrointestinal Toxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Mouse/ 104 wk 0, 0.03, 0.14, Stomach ulceration in males at 0.69 (p<0.05) Mitsumori et al.

B6C3F /60 M, ad lib in feed 0.69 (M); 0.03, BML not reported 1990 1

60 F 0.13, 0.6 (F)

(MMC)

3.3.3.6 Immunological

Suppression of the humoral and cellular immune responses have been observed in animals after

oral exposure to methylmercury or methylmercuric chloride. Both decreases in the production of

antibody-producing cells and decreased antibody titre in response to inoculation with immunestimulating agents (such as sheep red blood cells) have been observed (Blakley et al. 1980; Koller et al.

1977; Ohi et al. 1976). Decreases in natural killer T-cell activity have been observed in animals after

exposure to methylmercury (Ilback 1991).

Table 3-64

Immunotoxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex per Exposure Dose

Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/Brown Norway/6 NS x/wk 0, 4.8 IgG deposits along the glomerular capillary wall of Bernaudin et al 1981

both sexes exposed/22 2 mo (MMC) the kidney, not in arteries, suggestive of an

both sexes/controls autoimmune disease; no effect seen in controls.

Limitation: only one level tested

BML not reported

Mouse/ICR/6 M 5 d 0.27, 2.7 Decreased production of antibody-producing cells Ohi et al. 1976

1 x/d (MMC) (LOAEL = 2.7; p<0.01).

(gavage) Limitation: small number of animals, only males

tested

BML not reported

Mouse/Swiss/8-10 M 3 wk 0.076, 0.3, 1.52 Decreased production of antibody-producing cells Blakley et al. 1980

ad lib in (MMC) and decreased antibody titer (LOAEL = 0.076;

drinking water p<0.01).

Limitation: small number of animals, only males

tested

BML not reported
Table 3-64 (continued)

Immunotoxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex per Exposure Dose

Group Duration (mg/kg-day) Effects/Limitations/BML Reference

3-85

Mouse/Balb/c CUM/ 12 wk 0, 0.5 Reduced natural killer T-cell activity; decreased Ilback 1991

8 F ad lib in feed thymus weight and cell number (p<0.01).

Limitation: small number of animals treated, only

females tested

BML not reported

Rabbit/New Zealand 14 wk 0.04, 0.4, 0.8 Decreased antibody titer (LOAEL = 0.4) (26% of Koller et al. 1977

white/10 M, 10 F 1 x/d (MMC) the animals at 0.4 and no controls died by wk 14).

in feed Limitations: No statistical analysis

BML: 2,240 μg/L in blood at 0.4 mg/kg/d at wk

14

3.3.3.7 Dermal

Al-Mufti et al. (1976) studied the effects of methylmercury in humans who ate contaminated

bread; a correlation between bread consumption and a history of rash was reported. No other information

was located regarding dermal effects of organic mercury following oral exposure. 

Table 3-65

Dermal Toxicity of Methylmercury in Humans: Epidemiological Study

Species/ Exposure Dose

No. per Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/415 1-3 mo NS "History of rash" in 14% of exposed group, compared with Al-Mufti et al. 1976

exposed/1012 (feed) (MMC) <1% of unexposed

controls (sex Limitations: Effects poorly described; no statistical

NS) analysis

BML not reported

3.3.3.8 Developmental

Methylmercury readily crosses the placental barrier, and marked developmental toxicity has been

observed in both humans and animals after gestational exposures. Infants exposed to methylmercury

through the mother's milk or during gestation had elevated blood mercury levels, as did their mothers

(Amin-Zaki et al. 1976). Human data from epidemic poisonings that occurred in Japan (Harada 1978)

and Iraq (Amin-Zaki et al. 1974), as well as isolated exposures (Snyder and Seelinger 1976) indicate that

methylmercury predominantly affects the developing nervous system. Infants born to mothers who

ingested fish contaminated with methylmercury from Minamata Bay in Japan between 1953 and 1960
3-86

appeared normal at birth but within several months exhibited mental retardation, retention of primitive

reflexes, cerebellar symptoms, dysarthria, hyperkinesia, hypersalivation, strabismus and pyramidal

symptoms (Harada 1978). Similarly, infants born to mothers who had ingested bread made with seed

grain treated with methylmercury-containing fungicides in Iraq during 1971 to 1972 exhibited symptoms

ranging from delays in speech and motor development to mental retardation, reflex abnormalities and

seizures (Amin-Zaki et al. 1974, 1978). Histopathologic analyses of brain tissues from infants that died

in the Iraqi (Choi et al. 1978) and Minamata (Harada 1978) episodes showed atrophy and hypoplasia of

the cerebral cortex, corpus callosum and granule cell layer of the cerebellum; dysmyelination of the

pyramidal tracts; and/or abnormal neuronal cytoarchitecture characterized by ectopic cells and

disorganization of cellular layers.

A number of studies have attempted to evaluate developmental neurotoxicity in populations with

elevated methylmercury exposure from consumption of fish as a major component of the diet but for

whom massive poisonings have not been reported. Kjellstrom et al. 1989) observed a higher incidence of

abnormal scoring on tests designed to assess intelligence and development among children from New

Zealand whose mothers had high levels of hair mercury. Also a study by McKeown-Eyssen et al. (1983)

of a Cree population from northern Quebec revealed a correlation between maternal exposure (as

determined using hair levels) and abnormal muscle tone or reflexes in male children. A dose-response

for this effect was not observed. 

Dose-response analyses of human data from the Iraqi epidemic of 1971 to 1972 have indicated

correlations between maximal maternal hair levels during pregnancy and the severity of the neurological

deficits seen in the children (Cox et al. 1989; Marsh et al. 1981, 1987). An evaluation of a calculated

threshold for response is presented in Section 6.3.1 of this volume.

Table 3-66

Developmental Toxicity of Methylmercury in Humans: Case Studies

Species/ Exposure Dose

No. per Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/8 M, 7 ~2 mo. NS Assessment of 15 mother-infant pairs where the mothers Amin-Zaki et al.

F infants (feed) ate grain treated with methylmercury fungicide during 1974

pregnancy. Motor and mental development were impaired

(blindness, impaired hearing) in 6 infants; there were no

congenital malformations.

BML: Affected infants: ~3,000 μg/L in blood at 2 months;

Affected mothers: 400 μg/L in blood

Human/1 F 6 mo. NS Severe neurological impairment (blindness, myoclonic Snyder and

3 mo. seizures, spastic quadriparesis) of male infant born to a Seelinger 1976

postcoital- mother eating meat from pigs that had eaten grain treated

term with methylmercury fungicide.

(feed) Limitation: Case report

BML not reported
3-87

Table 3-67

Developmental Toxicity of Methylmercury in Humans: Epidemiologic Studies

Species/ Exposure Dose

No. per Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/220 F NS NS Mental retardation, atrophy of brain and degeneration of Harada 1978

(food) cerebellum in offspring. Of 220 infants born in Minamata

(to mothers eating contaminated fish), 13 had severe

symptoms; the number with less severe symptoms was not

reported.

Limitations: Few details on methods or results

BML not reported

Human/84 few days to NS Assessment of mother-infant pairs where mothers ate grain Marsh et al. 1981

mother-child several mo. treated with methylmercury fungicide during pregnancy

pairs (food) (same Iraqi population as reported by Amin-Zaki et al.

1974). Severe psychomotor retardation in infants.

BML Range: 37-293 μg/g in hair (maximum in segment of

maternal hair)

Human/243 Gestation and NS Abnormal tendon reflexes or muscle tone in male offspring McKeown-Eyssen et

exposed (sex lactation correlated with methylmercury exposure (p<0.05). al. 1983

NS) aged 12-30 (food) Conducted as a case-control study after potential affected

mo. measures were identified.

Limitation: Author reported that the statistical method

could have led to an association by chance.

BML avg: 6 μg/g in maternal hair 

Human/81 few days to NS Assessment of mother-infant pairs where mothers ate grain Marsh et al. 1987

mother-child several mo. treated with methylmercury fungicide during pregnancy

pairs (food) (same Iraqi population as reported by Amin-Zaki et al.

1974). Delayed walking and talking; seizures; mental

retardation. 

BML Range: ~18-598 μg/g (maximum in strand) in hair

of mothers of affected infants

The developmental toxicity of oral exposure to methylmercury has been extensively studied in

animals. In rodents exposed in utero, a spectrum of effects has been observed ranging from decreases in

fetal weight and skeletal ossification and increases in skeletal variations and malformations (brain

lesions, hydrocephalus, cleft palate, micrognathia, edema, subcutaneous bleeding, hydronephrosis,

hypoplasia of the kidneys, dilation of the renal pelvis) to increased resorptions and fetal deaths (Fuyuta et

al. 1978, 1979; Inouye and Kajiwara 1988a; Inouye and Murakami 1975; Khera and Tabacova 1973;

Nolen et al. 1972; Reuhl et al. 1981; Yasuda et al. 1985). The severity of the effects generally increased

with dose, and the incidence of malformations increased with exposures that occurred later in gestation

(Fuyuta et al. 1978; Inouye and Murakami 1975). Brain lesions have been observed in a variety of areas

including the brain mantle, corpus callosum, caudate putamen and cerebellum. In guinea pigs, early

gestational exposures (weeks 3–5 of pregnancy) resulted primarily in developmental disturbances of the

brain (smaller brains, dilated lateral ventricles and reduced size of caudate putamen), whereas later

gestational exposures (>week 6 of pregnancy) resulted in widespread neuronal degeneration (Inouye and

Kajiwara 1988b).

In addition to structural changes, functional changes have been observed in animals after

gestational exposures. Such functional effects include abnormal tail position during walking; flexion;
3-88

hindlimb crossing; decreased locomotor activity, responding in an avoidance task and righting response;

increased passiveness, startle-response and sensitivity to pentylenetetrazol-induced convulsions; and

impaired maze performance, operant behavior, swimming behavior, tactile-kinesthetic function, visual

recognition memory, temporal discrimination, and subtle learning deficits such as insensitivity to

changing reinforcement contingencies (Bornhausen et al. 1980; Buelke-Sam et al. 1985; Burbacher et al.

1990; Elsner 1991; Geyer et al. 1985; Gunderson et al. 1988; Hughes and Annau 1976; Inouye et al.

1985; Musch et al. 1978; Olson and Boush 1975; Rice 1992; Rice and Gilbert 1990; StoltenburgDidinger and Markwort 1990; Suter and Schon 1986; Newland et al. 1994).

Overt neurological impairment is the endpoint used to document methylmercury poisonings;

however, as shown in animal studies, methylmercury may produce more subtle neurodevelopmental

effects such as impairment of sensory or cognitive systems. Schreiner et al. (1986) exposed rats to 0, 0.2

or 0.6 mg Hg/kg-day as methylmercuric chloride in utero and during lactation to evaluate pup

performance on visual discrimination reversal task. While no overt signs of neurotoxicity were evident,

subtle differences between the control and high-dose group were observed during more difficult tasks. A

stressful or highly demanding situation appears to be necessary for the expression of these sensory

effects, wherein the decreased ability to adapt to the altered conditions became manifest. Spyker et al.

(1972) reported that although no signs of neurological toxicity was observed in mouse pups exposed to

methylmercury in utero, open field and swimming tests revealed subtle neurological effects in the

exposed pups. Newland et al. (1994) administered methylmercury by gavage to pregnant squirrel

monkeys between weeks 11 and 14.5 of gestation. Doses were adjusted to maintain 0.7 to 0.9 ppm Hg in

the maternal blood. There were three controls and three methylmercury-treated offspring. Offspring

were evaluated at 5-6 on a lever pressing test which required discrimination between degrees of

reinforcement. At steady state, monkeys exposed to methylmercury in utero were less sensitive to

differences in reinforcement rates. When reinforcement rates changed, exposed animals either changed

their behavior slowly in response to the altered reinforcement or not at all.

The developmental toxicity of methylmercury may be attributable to the ability of

methylmercury to bind to sulfhydryl-rich tubulin (a protein component of microtubules) and cause its

depolymerization (Falconer et al. 1994; Sager et al. 1983). Both cell division and cell migration require

intact microtubules for normal functioning. Disruption of microtubule function could result in the

derangement of cell migration (Choi et al. 1978; Falconer et al. 1994; Matsumoto et al. 1965) and

arrested cell division (Reuhl et al. 1994; Sager et al. 1984).

Table 3-68

Developmental Toxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/Charles 9 d 0, 0.02, 0.2, 4 Increased number of fetuses with soft tissue variations of Nolen et al. 1972

River/20 F Gd 6-14 the urinary system and incomplete ossification or

ad lib in calcification (LOAEL = 4; p<0.05).

drinking BML not reported

water
Table 3-68 (continued)

Developmental Toxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

3-89

Rat/Wistar/35 F 52 d 0, 0.002, 0.01, Increased incidence of eye defects (in harderian and Khera and Tabacova

ad lib in feed 0.05, 0.25 lachrymal glands) and salivary glands in fetuses (LOAEL 1973

(MMC) = 0.25); significant dose response (p = 0.01). Mothers

were treated from immaturity through weaning or later.

Limitations: Incomplete reporting; of results

BML not reported

Rat/Wistar/10 F 8, 12, or 20 d 2, 4 Increased brain lesions and generalized edema Inouye and

1 x/d (MMC) (Gd 0-20) (LOAEL = 2). Murakami 1975

Gd 12-20, Limitations: Limited data reporting; no statistical

0-12, or 0-20 analysis; small number of treated animals

(gavage) BML not reported

Rat/Holtzman/5 during gesta- 0, 2.5 Decreased visual evoked potential latencies for peaks N1 Zenick 1976

F tion, during (MMC) (p0.05), P1 (p0.01) and P2 (p0.01) in 30-day old

lactation, or pups exposed during gestation, during lactation, or

postnatal during postnatal days 21-30.

days 21-30 in BML not reported

drinking

water

Rat/Charles 47 d prior to 0.42, 0.7, 1.4 Ultrastructural changes, dose-related decrease in Fowler and Woods

River CD/20 F and during (MMH) biochemical activity in mitochondria of fetal hepatocytes 1977

gestation (p<0.01) following administration of methylmercury

ad lib in hydroxide to mothers (LOAEL = 1.4).

drinking BML: 40 μg/g (organic and inorganic) in liver of fetuses

water at 1.4 mg/kg-day

Rat/Long- Once 0, 4 Increased P1-N1 amplitudes and decreased P2 and N2 Dyer et al. 1978

Evans/4 exposed, Gd 7 (MMC) latencies of cortically visual evoked potential (p<0.05).

6 control (gavage) BML not reported

Rat/Wistar/20 F 8 d 0, 2, 4, 6 At 6 mg/kg-day, decreased maternal weight gain, Fuyuta et al. 1978

1 x/d (MMC) increased resorptions and fetal deaths (p<0.001);

Gd 7-14 decreased fetal body weight increased skeletal and

(gavage) visceral malformations (hydrocephaly, wavy ribs).

(LOAEL = 4; p<0.01)

BML not reported

Rat/Wistar- 4 d 0, 0.04, 1.6 Impaired ability to perform operant conditioning Musch et al. 1978

Neuherberg/ Gd 6-9 (MMC) procedures (number of responses on lever required in

No. F. NS (gavage) specified period of time) (LOAEL = 0.05).

Limitation: Statistical analyses not reported 

BML not reported

Rat/Wistar/10 F 4 d 0, 0.004, 0.008, Reduction in behavioral performance in offspring of Bornhausen et al.

1 x/d 0.035 treated mice following operant conditioning 1980

Gd 6-9 (MMC) (LOAEL = 0.008; p<0.01).

(gavage) BML not reported
Table 3-68 (continued)

Developmental Toxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

3-90

Rat/Sprague- Once 0, 6.3 Shorter avoidance latency in 60-day old offspring Cuomo et al. 1984

Dawley/No. F NS Gd 8 (MMC) (LOAEL = 6.3).

(gavage) BML not reported

Rat/Sprague- 10 d 0, 0.2, 1, 2, 4 Delayed sexual development (vaginal patency and testes Geyer et al. 1985

Dawley/No. F NS 1 x/d (MMC) descent), reduced pivoting, delayed surface righting,

Gd 6-15 partially retarded swimming development, increased

(gavage) activity in center of open field, impaired startle reflex

response. Reduced maternal weight gain and litter

weight. No live offspring were produced at 4 mg/kg-day

(LOAEL = 2; p<0.05).

BML not reported

Rat/Sprague- 4 d 0, 1.6, 4.8 Delayed vaginal patency, delayed surface righting, Vorhees 1985

Dawley/ 1 x/d (MMC) retarded swimming development, lower activity,

15-19 F Gd 6-9 impaired complex water maze performance. Increased

(gavage) mortality of pups at 1-21 days of age (LOAEL = 4.8;

p<0.05).

BML not reported

Rat/Wistar/ during 0, 0.2, 0.6 Increase in response latency in male (p<0.05) and female Schreiner et al. 1986

38 M, 38 F gestation and (MMC) pups (p<0.01) and in passiveness (p<0.05) in visual

lactation ad discrimination reversal task at 0.6 mg/kg-day (LOAEL =

lib in 0.6).

drinking BML not reported

water

Rat/HAN- 13 days prior 0, 0.2, 0.6, 1.7 Reduced weight gain, ataxia and inability to give birth in Suter and Schon

Wistar/10 F to mating (MMC) dams at 1.7. High mortality in pups at 1.7. Impaired 1986

until post- swimming behavior and righting reflex, delayed sexual

natal day 21 maturity (vaginal opening and testes descent) at 0.2 and

in drinking 0.6. (LOAEL = 0.2; p0.05).

water BML = 9,700-191,000 μg/L in dams and 10,000-

127,000 μg/L in pups at birth

Rat/Wistar/No. F 4 d 0, 0.02, 0.04, Increased startle response; impaired swimming behavior, StoltenburgNS 1 x/d 0.4, 4 decreased locomotor and nose-poking behavior; Didinger and

Gd 6-9 (MMC) alteration of dendritic spine morphology (LOAEL = 4). Markwort 1990

(gavage) Limitations: Limited data reporting; no statistical

analysis

BML not reported

Rat/Wistar/16 F 2 wk prior to 0, 0.08-0.38, Impaired tactile-kinesthetic function (p0.05) (LOAEL = Elsner 1991

mating 0.34-0.95 0.08-0.38).

through (MMC) BML not reported

weaning

ad lib in

drinking

water
Table 3-68 (continued)

Developmental Toxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

3-91

Rat/Sprague- Once 0, 6.4 Increased GABA receptors in prenatally exposed pups Guidetti et al. 1992

Dawley/No. and Gd 15 (MMC) sacrificed at 14 or 21 days postpartum; increased

sex NS (gavage) behavioral depression after diazepam.

A

Limitations: Only one treatment level; no data on

number of animals

BML not reported

Mouse/SvSl/ Once 0, Impaired swimming ability and open-field behavior Spyker et al. 1972

No. F NS Gd 7 or 9 0.16 mg (p<0.05) in 30-day old pups. Dose administered as

(i.p.) MMD/20 g methylmercury dicyandiamide (MMD)

BML not reported

Mouse/CFW/No. Once 0, 1, 2, 3, 5, 10 Increased number of trials to criterion (p<0.05) and Hughes and Annau

F NS Gd 8 (MMH) increased number that failed to attain criterion in 2-way 1976

(i.v.) avoidance test conducted on 56-day old pups (LOAEL =

3).

BML not reported

Mouse/ Once 0, 5, 7, 10 Longer center square latency at 10 (once) and 3.5 (3 d), Su and Okita 1976

129/Svsl/ Gd 10 decreased rearings and increased backings at 3.5;

No. F NS (s.c.) decreased locomotor activity at 7 and 10; postnatal

growth retardation at 7 and 10 (LOAEL = 7; p<0.05).

BML not reported

Mouse/C57BL/ 8 d 0, 2, 4, 4.8, 6 Increased resorptions and fetal deaths at 4.8 and 6 Fuyuta et al. 1978

10 F 1 x/d (MMC) (p<0.01); increased malformations (cleft palate, fused

Gd 6-13 vertebrae) at 2 and higher (p<0.05); increased skeletal

(gavage) variations; decreased maternal weight gain at 4.8 mg/kgday (LOAEL = 2).

Limitation: small number of treated animals

BML not reported

Mouse/ Once 0, 8 Arrest of brain cells during mitosis (p<0.01). Rodier et al. 1984

DUB/ICR/8 F Gd 12 Limitations: Only one dose tested; small number of

exposed, 7 F (gavage) animals tested

controls BML not reported

Mouse/ Once 0, 16 Decreased neonatal survival and weight gain; impaired Inouye et al. 1985

C3H/HeN/10 F Gd 13, 14, (MMC) righting response; decreased locomotor activity;

15, 16, or 17 abnormal gait; crossing of hindlimbs; decreased brain

(gavage) weight in groups treated on Gd 13 or 14 (p<0.01); dilated

lateral ventricles; slightly simplified cerebellar pattern. 

Effects were seen in groups dosed on all days, but

somewhat stronger in those treated on Gd 13 or 14.

Limitations: Incomplete reporting of data; most

parameters were not analyzed statistically; only one dose

tested

BML: ~20 μg/g in brain of fetuses
Table 3-68 (continued)

Developmental Toxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

3-92

Guinea pig/ Once 9.4-15 Aborted litters and retarded fetal brain development at all Inouye and Kajiwara

Hartley/5-9 F Gd 21, 28, 7.5 mg/animal treatment times. 1988b

35, 42, or 49 (wt 500-800 g) Limitations: No statistical analysis; small number of

(gavage) (MMC) treated animals, only 1 day of dosing

Avg BML over treatment time: Fetal: 2,600 μ/L in

blood; Maternal: 1,800 μg/g in blood

Hamster/ Once at 0, 1.6, 8 Degeneration of cerebellar neurons in rats born to Reuhl et al. 1981

Golden/10 F Gd 10, or (MMC) mothers treated with 1.6 mg/kg/d on Gd 10-15 or a

6 d single dose of 8 mg/kg on Gd 10 and sacrificed

1 x/d neonatally or as adults.

Gd 10-15 Limitation: small number of treated animals

(gavage) BML not reported

Monkey/Macaca approx. 1-3 0, 0.04, 0.06 Impaired visual recognition memory (data pooled from Gunderson et al.

fascicularis/9 F yr 1 x/d prior both groups of infants of exposed mothers) compared to 1988

exposed, 8 F to mating unexposed controls; test performed at 50-60 days of age.

control through Limitation: small number of treatment animals

gestation (in BML Range: 880-2,450 μg/L in blood of infants at birth;

apple juice) 280-830 μg/L at testing

Monkey/Macaca approx. 4 mo 0, 0.04 Decrease in social play behavior and concomitant Gunderson et al.

fascicularis/ to 2 yr 1 x/d increase in nonsocial passive behavior compared to 1988

12 F exposed, 13 prior to unexposed controls; tests performed at 2 weeks to

F control mating 8 months of age.

through Limitation: small number of treatment animals

gestation (in BML Range: 1,565 μg/L in blood of infants at birth

apple juice)

Monkey/Macaca 4-4.5 yr 0, 0.01, 0.025, Spatial visual impairment (LOAEL = 0.01). Rice and Gilbert

fascicularis/5 1 x/d in utero 0.5 Limitation: Small number of infants (5 high-dose; 2 1990

mothers and (MMC) mid-dose; 1 low-dose)

postnatally BML not reported

(gavage)

Monkey/ 6.5-7 yr 0, 0.05 Six years after end of dosing (follow-up study to Rice Rice 1989b; Rice

Macaca 7 d/wk (MMC) and Gilbert 1982); decreased fine motor performance; and Gilbert 1992

fascicularis/4 M, 1 x/d diminished touch and pinprick sensitivity; impaired high

1 F exposed, 1 (capsule; frequency hearing (p<0.05).

M, 2 F controls gavage) Limitations: small number of animals tested; one dose

level tested

BML: Not detectable at time of dosing

Monkey/Macaca 4-4.5 yr 1xd 0, 0.01, 0.025, Monkeys tested as juveniles showed no gross intellectual Rice 1992.

fasicularis/13 in utero and or 0.05 impairment; some indication of decreased temporal

total postnatally discrimination. BML in treated animals at birth

(gavage) averaged 0.46, 0.93, or 2.66 ppm; decreased to steadystate of 0.20, 0.25 or 0.60 ppm.
Table 3-68 (continued)

Developmental Toxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

3-93

Monkey/Macaca unspecified 0.04, 0.06, 0.08 No effect on spatial memory of adult offspring of Gilbert et al. 1993

fascicularis/23 F period prior animals treated with methylmercury hydroxide (data

to mating pooled from 24 animals, all treated groups).

through BML Range: 1,040-2,460 μg/L in blood of infants at

gestation birth

Monkey/Saimiri week 11 or 0.7 to 0.9 ppm Monkeys exposed in utero tested (on learned lever Newland et al. 1994

sciureus/3 F 14.5 until methylmercury pulling activity) at ages 5-6 yr. Methylmercury

parturition in maternal treatment resulted in decreased sensitivity to degrees in

(gavage) blood reinforcement; change in reinforcement degree resulted

in either no behavior change or slow change by

comparison to controls. Limitations: small number of

animals tested; incomplete reporting on treatment.

3.3.3.9 Reproductive

Although no data were located regarding the reproductive effects of oral exposure to

methylmercury in humans, animal data suggest that, at sufficiently high doses, methylmercury may

adversely affect reproductive function in both males and females. When male rats were given

methylmercury for several days prior to mating, mated females were observed with increased

preimplantation losses (Khera 1973). Exposure of male monkeys to methylmercury for longer durations

has been shown to adversely affect sperm motility and speed and to result in increased incidences of

sperm tail defects (Mohamed et al. 1987). Decreases in spermatogenesis and tubular atrophy of the testes

have been observed upon histopathological analyses of the testes of mice exposed to methylmercury

chronically (Hirano et al. 1986; Mitsumori et al. 1990).

Less information is available regarding the effects of methylmercury on female reproductive

function. Exposure of female monkeys to methylmercury for 4 months prior to mating produced no

effects on the length of the menstrual cycle but resulted in decreased conceptions and increased early

abortions and stillbirths (Burbacher et al. 1988). Several studies have shown increased rates of

resorptions and abortions after exposure during gestation (Fuyuta et al. 1978; Hughes and Annau 1976;

Inouye and Kajiwara 1988a); however, it is unclear from these studies whether the effects observed are

the result of maternal reproductive failure or fetal toxicity.
3-94

Table 3-69

Reproductive Toxicity of Methylmercury in Animals

Species/

Strain/ Exposure Dose

No. per Sex per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Rat/Wistar/10-20 M 7 d 0, 1, 2.5, 5 Reduced mean litter size after male exposure Khera 1973

1 x/d (MMC) (LOAEL = 5; p<0.01) in sequential mating trials

(gavage) with unexposed females

BML not reported

Rat/Wistar/14-19 M 95-125 d 0.1, 0.5, 1 Males were mated to unexposed females concurrent Khera 1973

1 x/d (MMC) with dosing. Reduced mean litter size (LOAEL =

0.5)

BML not reported

Mouse/Swiss Webster/10-20 5-7 d 0, 1, 2.5, 5 No effect on number of viable embryos, dead Khera 1973

M 1 x/d (MMC) embryos, or percent pregnancy (NOAEL = 5)

(gavage) BML not reported

Mouse/ICR/60 M, 60 F 104 wk 0, 0.03, 0.15, Significantly decreased spermatogenesis (LOAEL = Hirano et al.

ad lib in feed 0.72 (M); 0.02, 0.73; significance level not reported) 1986

0.11, 0.62 (F) BML not reported

(MMC)

Mouse/B6C3F /60 M, 60 F 104 wk 0, 0.03, 0.14, Tubular atrophy of the testes (LOAEL = 0.69; Mitsumori et 1

ad lib in feed 0.68 (M); 0.03, p<0.01) al. 1990

0.13, BML not reported

0.6 (F)

(MMC)

Monkey/Macaca 20 wk 0, 0.047, 0.065 Decreased sperm motility and speed; increased Mohamed et

fascicularis/3 M 7 d/wk sperm tail defects (LOAEL = 0.065; p<0.05) al. 1987

1 x/d BML: ~2200 μg/L in blood at 0.065 mg/kg-day,

(gavage) approaching steady state

Monkey/Macaca 4 mo prior to 0, 0.04, 0.06, Abortion; stillbirth; decreased conception in exposed Burbacher et

fascicularis/7-9 F mating 0.08 females (LOAEL = 0.06); no effect on menstrual al. 1988

1 x/d (MMH) cyclicity

(gavage) Avg. BML: 1,600 μg/L in blood at equilibrium at

0.06 mg/kg

3.3.3.10 Genotoxicity

Data from several studies in humans suggest that ingesting methylmercury may cause

chromosomal aberrations and sister chromatid exchange (Skerfving et al. 1970, 1974; Wulf et al. 1986;

Franchi et al. 1994).

A study of nine Swedish subjects who consumed mercury-contaminated fish and 4 controls

showed a statistically significant rank correlation between blood mercury and percentage of lymphocytes

with chromosome breaks (Skerfving et al. 1970). An extension of this study (Skerfving et al. 1974)

included 23 "exposed" (5 females and 18 males) and 16 "controls" (3 females and 13 males). The authors
3-95

reported a significant correlation between blood mercury level and frequency of chromatid changes and

"unstable" chromosome aberrations; there was no correlation with "stable" chromosome aberrations.

The Wulf et al. (1988) study was of 92 Greenlander Eskimos. Subjects were divided into three

groups based on intake of seal meat (6 times/week; 2-5 times/week; once/week or no consumption of seal

meat). Higher frequency of SCE in lymphocytes was correlated with blood mercury concentration; an

increase of 10 μg Hg/L in blood was associated with an increase of 0.3 SCE/cell. Positive correlations

were also found for smoking, diet, living district and cadmium exposure.

Franchi et al. (1994) evaluated formation of micronuclei in peripheral blood lymphocytes of

Mediterranean fishers, a group with presumed high exposure to methylmercury. Fifty-one subjects were

interviewed on age, number of seafood-based meals/week and habits such as smoking and alcohol

consumption. Total blood mercury was measured; the range was 10.08 – 304.11 ng/g with a mean of

88.97 + 54.09 ng/g. There was a statistically significant correlation between blood mercury concentration

and micronucleus frequency and between age and micronucleus frequency.

Table 3-70

Genotoxicity of Methylmercury in Humans: Case Study

Species/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/6 M, 3 >5 yr NS Correlation between blood mercury concentration and Skerfving et al. 1970

F exposed; 3 3 x/wk chromosome breaks in lymphocytes cultured from people

M, 1 F control who ate mercury-contaminated fish

Limitation: Small sample size; limited exposure data

BML Range: 4-650 μg/L in blood

Table 3-71

Genotoxicity of Methylmercury in Humans: Epidemiology Study

Species/ Exposure Dose

No. per Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

Human/24-63 NS NS Incidence of sister chromatid exchanges (SCEs) in Wulf et al. 1986

(both sexes) cultured peripheral lymphocytes correlated with intake of

seal meat in an Eskimo population (as a surrogate for

mercury intake); p = 0.001. Other factors also correlated

with SCEs, but multiple regression analysis found that

some of the effect was attributable to mercury.

Limitation: Limited exposure data

BML not reported
Table 3-71 (continued)

Genotoxicity of Methylmercury in Humans: Epidemiology Study

Species/ Exposure Dose

No. per Sex Duration (mg/kg-day) Effects/Limitations/BML Reference

3-96

Human / 51 M measured as NS Incidence of micronuclei positively correlated with blood Franchi et al. 1994.

seafood mercury concentration and with age. No correlation with

meals/ week. smoking or number of seafood meals /week. Limitation:

Range 2 - 14. no control group.

BML range: 10.08 - 403.11 g/g blood.

Human/18M 10.5 yr 0.15-0.44 Increased frequency of chromosomal breaks. Popescu et al. 1979

exposed/10 (occup) (HgCl Limitations: Workers also exposed to mercuric chloride

control and one worker had history of benzene poisoning; control 2)

group was not matched for sex, smoking habits, or sample

size.

BML: 890 g/L in urine (avg)

In a study with cats (Charbonneau et al. 1976), methylmercury did not induce dose-related

unscheduled DNA synthesis in lymphocytes or chromosomal aberrations in bone marrow cells after oral

exposure to methylmercury for up to 39 months (Miller et al. 1979). Statistically significant decreases in

unscheduled DNA synthesis and increases in chromosomal aberrations were observed, but there was no

dose-response.

Table 3-72

Genotoxicity of Methylmercury in Cats

Species/

Strain/

No. per Sex Exposure Dose

per Group Duration (mg/kg-day) Effects/Limitations/BML Reference

Cat/Breed and 39 mo 0.008, 0.020, No dose-related changes in unscheduled DNA synthesis in Miller et al. 1979

sex NS 7 d/wk 0.046 cultured lymphocytes or frequency of chromosomal

aberrations in bone marrow of cats fed mercurycontaminated fish or a fish diet supplemented with

methylmercuric chloride

Limitations: No positive control; no assessment of

cytotoxicity

BML Range: 500-13,500 μg/L Hg in blood

Strain-specific differences exist with respect to the ability of methylmercury to produce dominant

lethal effects in mice (Suter 1975). When (SEC x C Bl)F males were injected with 10 mg/kg 57 1

methylmercury hydroxide, there was a slight reduction in the total number of implantations and a

decrease in the number of viable embryos. This was not observed when (101 x C H)F males were 3 1
3-97

exposed in a similar fashion. When female (10 x C H)F mice were treated with methylmercuric 3 1

hydroxide, no increase in the incidence of dead implants was observed (unlike the case for mercuric

chloride). Changes in chromosome number but no increase in chromosome aberrations were observed in

oocytes of Syrian hamsters treated with one i.p injection of 10 mg/kg methylmercuric chloride (Mailhes

1983). Methylmercury was administered s.c. to golden hamsters at doses of 6.4 mg or 12.8 mg

Hg/kg/body weight. Polyploidy and chromosomal aberrations were increased in bone marrow cells, but

there was no effect on metaphase II oocytes. There was an inhibitory effect on ovulation which the

authors noted was not as severe as that induced by mercuric chloride in the same study (Watanabe et al.

1982). Non-dysjunction and sex-linked recessive lethal mutations were seen in Drosophila melanogaster

treated with methylmercury in the diet (Ramel 1972).

As reviewed in WHO (1990), methylmercury is not a potent mutagen but is capable of causing

chromosome damage in a variety of systems. In vitro studies have generally shown clastogenic activity

but only weak mutagenic activity. Methylmercuric chloride and dimethylmercury were both shown to

induce chromosome aberrations and aneuploidy in primary cultures of human lymphocytes;

methylmercuric chloride was the more potent clastogen at equally toxic doses (Betti et al. 1992). Both

methylmercury and mercuric chloride induced a dose dependent increase in SCE in primary human

lymphocytes and muntjac fibroblasts; methylmercury was about five time more effective in this regard

(Verschaeve et al. 1984; Morimoto et al. 1982).

Methylmercury has been shown to inhibit nucleolus organizing activity in human lymphocytes

(Verschaeve et al. 1983). Methylmercury can induce histone protein perturbations and has been reported

to interfere with gene expression in cultures of glioma cells (WHO 1990). Impaired growth and

development was noted in cultured mouse embryonic tissue treated in vitro with methylmercuric

chloride, but there was no increase in SCE (Matsumoto and Spindle 1982). Costa et al. (1991) showed

that methylmercuric chloride caused DNA strand breaks in both V79 and rat glioblastoma cells treated in

vitro. Methylmercuric chloride produced more strand breaks than did mercuric chloride.

Evidence of DNA damage has been observed in the Bacillus subtilis rec-assay (Kanematsu et al.

1980). These authors reported negative results for methylmercury in spot tests for mutagenicity in the

following bacterial strains: E. coli B/r WP2 and WP2; and Salmonella typhimurium strains TA1535,

TA1537, TA1538, TA98 and TA100. Jenssen and Ramel (1980) in a review article indicated that

methylmercury acetate was negative in both micronucleus assays and in mutagenicity tests in Salmonella;

the article referred to Heddle, J.R. and W.R. Bruce (1977) and provided no experimental details. Weak

mutagenic responses for methylmercuric chloride and methoxyethyl mercury chloride were observed in

Chinese hamster V79 cells at doses near the cytotoxic threshold (Fiskesjo 1979), and methylmercury

produced a slight increase in the frequency of chromosomal nondisjunction in Saccharomyces cerevisiae

(Nakai and Machida 1973). Methylmercury, however, caused neither gene mutations nor recombination

in S. cerevisiae (Nakai and Machida 1973). Methylmercury retarded DNA synthesis and produced single

strand breaks in DNA in L5178Y cells (Nakazawa et al. 1975).
4-1

4. SUSCEPTIBLE POPULATIONS 

A susceptible population is a group who may experience more severe adverse effects at

comparable levels or adverse effects at lower exposure levels than the general population. The greater

response of these sensitive subpopulations may be a result of a variety of intrinsic or extrinsic factors. 

Volume V describes populations that may be at increase risk because of higher exposure to mercury and

mercury compounds. Additional factors that may be important include, but are not limited to, the

following: an impaired ability of the detoxification, excretory, or compensatory processes in the body to

protect against or reduce toxicity; differences in physiological protective mechanisms (e.g., blood brain

barrier); or unique toxic reactions that are specific to the genetic makeup, developmental stage, health

status, gender or age of the individual.

The nervous and renal systems are the primary targets for mercury-induced toxicity. Data are

also available indicating some effects to the respiratory, cardiovascular, gastrointestinal, hematologic,

immune, and reproductive systems. The developing organism appears to be particularly sensitive to

methylmercury exposure. In addition, it is probable that individuals with preexisting damage or disease

in target organs for mercury-induced toxicity may experience more severe effects upon exposure to

mercury. The populations listed below may be highly susceptible to mercury toxicity.

 Developing Organisms. Data from epidemic poisonings in Japan (Harada 1978) and Iraq

(Marsh et al. 1987) indicate that infants exposed in utero to methylmercury developed

marked neurological development delays while their mothers experienced little or no

overt signs of toxicity. Data indicate that the developing fetus may be 5 to 10 times

more sensitive than the adult (Clarkson, 1992). This difference in sensitivity is believed

to be due, in part, to the high sensitivity of developmental processes (i.e., cellular

division, differentiation, and migration) to disruption by mercury (Choi et al. 1978;

Sager et al. 1982). One factor that may account for this difference in sensitivity is the

presence of an incomplete blood brain barrier in the fetus. Another important factor may

be the lack of methylmercury excretion in the fetus (Grandjean et al. 1994).

 Age - Infants and Other Age Groups. Available data indicate that neonates are at

increased risk to inorganic mercury and methylmercury. Both inorganic and organic

forms of mercury are excreted in breast milk (Sundberg and Oskarsson 1992; Yoshida et

al. 1992; Grandjean et al. 1994); thus, neonates in an exposed population may experience

increased mercury exposure. Animal data for rats indicate that suckling infants retain a

higher percentage of ingested inorganic mercury than do adults (Kostial et al. 1978). 

The most significant difference in organ retention (neonates > adults) was

methylmercury in the brain following exposure to methylmercury (Yang et al. 1973;

Kostial et al. 1978) and inorganic mercury retained in the kidney following exposure to

elemental mercury (Yoshida et al. 1992). These differences may be associated with an

increased absorption of mercury with a milk diet, a decrease in excretion, or an

incomplete blood brain barrier (Kostial et al. 1978, Grandjean et al. 1994).

Signs of toxicity may begin to be manifested several years after the cessation of dosing,

possibly related to subclinical effects being unmasked by aging. Rice (1989b) dosed

monkeys with methylmercury from birth to 6.5-7 years of age. Although there were no

overt signs of neurotoxicity during dosing, neurological deficits were observed at 13
4-2

years of age, 6-7 years following cessation of exposure. Similarly, a small human

population with Minamata disease has been identified in Japan as experiencing new or

worsening neurological effects a few years following termination of mercury exposure. 

This late-onset Minamata disease may be related to several factors including aging (Igata

1993).

 Gender. Sex-related differences in mercury toxicokinetics and sensitivity to mercury

have been observed, although data indicate that the more sensitive sex may differ by

species and strain. Using death as the critical endpoint, in one strain of mice,

C57BL/6N, males were less sensitive to methylmercury following daily dosing than

females while, in contrast, male mice were more sensitive than females in another strain,

BALB/cA (Yasutake and Hirayama 1988). In humans, although the ratio of males to

females with Minamata disease has been reported to be 1.2:1, the ratio of deaths was

recorded at 1.8:1 (Tamashiro et al. 1984).

Other studies are in general agreement that male rats (Thomas et al. 1986) and mice

(Nielsen and Andersen 1991a, 1991b) eliminate mercury faster and have lower tissue

levels than females following dosing with methylmercury. Part of the difference in

whole-body retention of mercury in methylmercury-exposed mice has been associated

with varying degrees of deposition of mercury in the carcass, including the skin and hair

(Nielsen and Andersen 1991b). This difference is thought to be due in part to

differences in glutathione metabolism and renal excretion of mercury, which is affected

by the hormonal status of testosterone (Nielsen et al. 1994). Hirayama et al. (1987)

have reported that the toxicokinetics of methylmercury in castrated male mice was very

similar to that in female mice, and that the male pattern of methylmercury toxicokinetics

could be restored by testosterone treatment. Such differences were not observed in a

small set of similarly tested human volunteers (Miettinen et al. 1971).

 Dietary Insufficiencies of Zinc, Glutathione, or Antioxidants. Mercury has been

suggested to cause tissue damage by increasing the formation of reactive oxygen species

and activation of lipoperoxidation, calcium-dependent proteolysis, endonuclease activity,

and phospholipid hydrolysis (Ali et al. 1992; LeBel et al. 1990, 1992; Gstraunthaler et

al. 1983; Verity and Sarafian 1991). Zinc, glutathione, and antioxidant deficiencies

would be expected to exacerbate mercury-induced damage by limiting cellular defenses

against the oxidative processes. Animal data support the importance of zinc,

glutathione, and antioxidants in limiting mercury-induced damage (Fukino et al. 1992;

Girardi and Elias 1991; Yamini and Sleight 1984) (see also Section 5, Interactions).

 Predisposition for Autoimmune Glomerulonephritis. Autoimmune glomerulonephritis is

a form of renal toxicity characterized by proteinuria, deposition of immune material (i.e.,

autoantibodies and complement C3) in the renal mesangium and glomerular blood

vessels and glomerular cell hyperplasia (Bigazzi 1992; Goldman et al. 1991; Mathieson

1992). Limited human data suggest that certain individuals may develop this

autoimmune response when exposed to inorganic or elemental mercury (Cardenas et al.

1993; Langworth et al. 1992b; Tubbs et al. 1982). While the etiology of this syndrome

has not been completely elucidated, data from susceptible and resistant strains of animals

indicate that susceptibility is governed by both major histocompatibility complex (MHC)
4-3

genes and non-MHC genes (Aten et al. 1991; Druet et al. 1978; Hultman and Enestrom

1992; Hultman et al. 1992; Michaelson et al. 1985; Sapin et al. 1984).

 Predisposition for Acrodynia. Acrodynia, also known as "pink disease," is a

hypersensitive response following exposure to elemental or inorganic mercury and is

characterized by the following signs and symptoms: irritability; marked mood swings;

restlessness; itching; flushing, swelling, and/or desquamation of the palms of the hands

and soles of the feet (the tip of the nose, ears, and cheeks may also be affected);

excessive perspiration; loss of appetite; tachycardia; hypertension; joint pains and

muscle weakness; photophobia; and sleeplessness. Acrodynia, which is more likely

related to exposure level rather than any inherent, genetic sensitivity, rarely occurs in the

general population.

Limited reports indicate that acrodynia has been almost exclusively observed in children,

affecting approximately 1 in 500 exposed children (Blondell and Knott 1993; Warkany

and Hubbard 1953). This disease was recently observed in a 4-year-old Michigan boy

who was exposed to mercury vapor released from paint in which mercury had been used

as a fungicide (Aronow et al. 1990). In this case, family members (i.e., both parents and

two siblings) were also exposed to the mercury vapors but remained asymptomatic

(Aronow et al. 1990). This case study supports the hypothesis that there is no genetic

predisposition to acrodynia.

Acrodynia was more frequently observed in the past when mercury-containing laxatives,

worming medications, teething powders and diaper rinses were widely used (Gotelli et

al. 1985; Warkany and Hubbard 1953). The physiological basis for this hypersensitivity

has not been identified. It does not appear, however, to be an allergic reaction to

mercury or to occur in the most highly exposed individuals (Warkany and Hubbard

1953).
5-1

5. INTERACTIONS

A number of interactions have been identified for chemicals that affect the pharmacokinetics

and/or toxicity of mercury compounds. Table 5-1 summarizes interactions in which potentiation or

protection from the toxic effects of mercury have been observed. Interactions that affect mercury

toxicokinetics are also shown. The effect on toxicity, however, can not be predicted based on changes in

distribution or excretion. For example, zinc pretreatment increases renal mercury levels but decreases

toxicity because it alters the distribution within the kidney (Zalups and Cherian 1992).

Only the interaction of selenium with mercury will be discussed in detail here. Selenium is

known to bioaccumulate in fish, so exposure to methylmercury in fish is associated with exposure to

increased levels of selenium. Where the main source of dietary mercury is fish, the diet is naturally

enriched with selenium relative to mercury. Increased selenium has been suspected of providing some

degree of protection, either by preventing oxidative damage or by forming a methylmercury-selenium

complex (Grandjean 1992a). It does not appear that the population in the Iraqi poisoning incident was

selenium deficient. Animal studies have demonstrated that simultaneous ingestion of selenium may be

protective against toxicity of methylmercury based upon its antioxidant properties (see Table 5-1). This

may explain why the latent period in Japan, where the population was exposed to methylmercury in fish,

was longer than that in Iraqi, where the exposure was to methylmercury in grain.

A common association between the metabolism of selenium and methylmercury is the thiolcontaining peptide glutathione (GSH). The metabolic cycling and oxidation-reduction of GSH are

integral processes coupled to the activation and metabolism of selenium (Hill and Burke 1982) and the

metabolism and detoxification of methylmercury (Ballatori and Clarkson 1982; Thomas and Smith

1982).

There are data to indicate that selenium co-administered with methylmercury can form seleniummethylmercury complexes (Magos et al 1987). The formation of these complexes appeared temporarily

to prevent methylmercury-induced tissue damage but also apparently delayed excretion of the

methylmercury in the urine. Thus, formation of selenium-methylmercury complexes may not reduce

methylmercury toxicity but may rather delay the onset of symptoms.

In support of the protective role of biological selenium, several investigators have found that a

diet supplemented with seafood high in selenium delayed the onset of methylmercury intoxication in rats

(Ganther 1980; Ohi et al. 1976). Ganther (1980) has observed that rats given selenium plus

methylmercury show increased body burdens of both selenium and methylmercury without signs of

toxicity. The accumulation of these elements may lead to mutual detoxification, but such

coaccumulation is not always linked to protection. Fair and associates (1985) have examined renal

ultrastructure changes along with changes in gamma glutamyl transferase activity in mice coadministered

both selenium and methylmercury in diet for 7 or 20 days or given a single i.p. dose. The results of this

study indicated that dietary selenium had only an initial protective effect against mercury accumulation

in the kidney; injected selenium offered longer protection.

Selenium has been shown to protect against the developmental toxicity of methylmercury in mice

(Nishikido et al. 1987; Satoh et al. 1985) and protects against oxidative damage by free radicals (CuvinAralar and Furness 1991; DiSimplicio et al. 1993; Ganther 1978). Further studies reported by

Fredricksson et al. (1993) indicated that dietary selenium supplementation during pregestation through

lactation in rats resulted in reduction of some adverse effects (hypoactivity) in neonates of the
5-2

methylmercury administered to mothers by gavage during organogenesis period. Significant increases in

glutathione peroxidase activity were noticed in animals fed selenium supplemented diet.

 Kosta et al. (1975) have observed a coaccumulation of mercury and selenium in the organs and

tissues of mineworkers at an approximate molar ratio of 1:1. In these circumstances, the abnormally high

mercury levels detected in the tissues were without apparent adverse effects on the miners. After

exposure to mercury in the mines, several of the miners had been retired 10–16 years when the study was

conducted. The selenium intake from the diet was not reported but was said not to be abnormally high,

suggesting that the co-accumulation with mercury is a natural and autoprotective effect. It is plausible

that in areas naturally low in selenium, individuals would be at greater risk from methylmercury

poisoning than those in areas of high selenium concentration. 

A group of 21 workers with no previous history of mercury exposure were monitored for urinary

mercury and selenium after their employment in the demolition of a chlor-alkali plant. Pre-exposure

urinary mercury ranged from 0.3 - 1.9 nmol/ mmol creatinine (mean = 0.8); urinary selenium was 13.9 -

89.5 (mean = 39.1). Post-exposure urinary mercury was significantly increased; 1.2 - 10.0 nmol/ mmol

creatinine (mean = 4.8). Selenium in the urine was decreased post exposure to 10.1 - 52.9 nmol/ mmol

creatinine (mean = 29.0). The authors did not speculate on the biological significance of the change in

urinary selenium. 

Co-administration of methylmercury and selenium apparently results in decreased

methylmercury concentrations in kidney; mercury levels in brain and liver, however, are increased

(Suzuki and Yamamoto, 1984; Brzenicka and Chmielnicka, 1985; Komsta-Szumska et al., 1983). 

Selenium has also been observed to increase methylmercury staining in spinal cord and nerve cell bodies

(Møller-Madsen and Danscher, 1991). A positive correlation between brain mercury and selenium 

levels was observed in monkeys exposed to methylmercury with no additional exposure to selenium other

than that in a standard diet (Björkman et al., 1995). The apparent protective effect of selenium against

overt high-dose methylmercury toxicity has been attributed to the decreased accumulation of

methylmercury in kidney in the presence of selenium (Stillings et al., 1974). It is doubtful that this effect

on kidney is relevant at environmental levels of methylmercury. It has also been suggested that the

formation of bis (methylmercury) selinide may render methylmercury less toxic (Naganuma and Imura,

1980), but there is no direct evidence for this. In addition, although increased fish consumption was

associated with a very modest increase in (cord) blood selenium levels in a fish-eating population, the

blood mercury levels increased much more dramatically (Grandjean et al., 1992). Grain may also

contain substantial levels of selenium, depending on the soil in which it is grown. Based on the

questionable relevance of any protective effect of selenium against high-dose methylmercury

nephrotoxicity, the fact that increased selenium intake results in increased brain mercury levels following

methylmercury ingestion, and a complete lack of data on the selenium status of the Iraqi population

exposed to methylmercury via grain, there is no reason to postulate that ingestion of methylmercury in a

fish matrix would result in decreased toxicity compared to ingestion in a non-fish matrix. Indeed, the

study in cats addressed this directly found no difference in toxicity or tissue levels when methylmercury

was administered as contaminated fish or added to a non-fish meal (Charbonneau et al., 1974).
5-3

Table 5-1

Interactions of Mercury with Other Compounds

Compounds Effects Observed Proposed Underlying Mechanism(s) References

Diethylmaleate and Increased renal toxicity Diethylmaleate causes depletion of nonprotein Girardi and Elias

inorganic mercury sulfhydryls 1991

Ethanol and Potentiated toxicity Unknown; increased mercury concentrations were Rumbeiha et al.

methylmercury observed in brain and kidneys, but changes in 1992

Increased mortality, mercury content were insufficient to fully explain Tamashiro et al.

increased severity of the potentiated toxicity 1986

neurotoxicity, renal Turner et al. 1981

toxicity

Decreased time to onset of

neurotoxicity

Ethanol and elemental No data on toxicity Inhibition of oxidation of metallic mercury to Nielsen-Kudsk 

mercury mercuric mercury by catalase 1965

Decreased mercury Magos and Webb 

absorption The effect on toxicity can not be predicted, due to 1979

Increased mercury levels Dencker 1982, 

in liver and in fetus 1984b

the opposing effects. Khayat and 

Ethanol and inorganic No data on toxicity Elemental mercury was exhaled, suggesting that Dunn et al. 1981b

mercury ethanol increased the activity of an unidentified

Increased mercury enzyme that reduces mercuric mercury to

exhalation elemental mercury.

Because elemental mercury, but not mercuric

mercury, can cross the blood brain barrier and the

placenta, toxicity to the brain and the developing

fetus may be increased

Thiol compounds Protection from renal Competition for protein binding sites; subsequent Magos and Webb

[e.g., N- toxicity increases in urinary excretion of mercury 1979

acetylpenicillamine,

penicillamine, 2-

mercapto propanol

(BAL)] and inorganic

mercury

Selenium and mercury Increased survival Mercuric mercury and selenium form a complex Parizek and

(simultaneous exposure) with a high molecular weight protein Ostadolva 1967

Decreased or delayed Satoh et al. 1985

renal, developmental Methylmercury forms a bismethylmercury Naganuma and

toxicity selenide complex Imura 1981

Potential mechanisms for protection: 1980

 -redistribution from sensitive targets Cuvin-Aralar and

 -competition of selenium for mercury Furness 1991

 binding sites associated with toxicity Imura and

 -increased selenium available for selenium- Naganuma 1991

 dependent glutathione peroxidase Nylander and

 (prevention of oxidative damage) Weiner 1991

Mengel and Karlog

Tellurium and elemental Decreased toxicity (effect Complexation of tellurium with mercury, by Magos and Webb 

or inorganic mercury unspecified) analogy to the chemically-related selenium 1979

Retention in body Khayat and 

increased Dencker 1984a
Table 5-1 (continued)

Interactions of Mercury with Other Compounds

Compounds Effects Observed Proposed Underlying Mechanism(s) References

5-4

Potassium dichromate Decreased renal function Unknown; both chemicals are toxic to the renal Baggett and Berndt

and inorganic mercury (measured as inhibition of proximal tubule 1984

p-aminohippurate

transport)

Zinc pre-treatment and Some protection from Zinc pretreatment induces metallothionein binding Zalups and

inorganic mercury nephrotoxicity of in kidneys Cherian 1992

inorganic mercury

Mercury binds preferentially to metallothionein, so

that less mercury is available to cause oxidative

damage in the proximal tubules

Zinc-deficiency and Exacerbation of renal Zinc-deficiency and mercury both independently Fukino et al. 1992

inorganic mercury toxicity increase renal oxidative stress

Together, the protective mechanisms of the kidney

are overwhelmed and oxidative damage is

compounded

Atrazine and Early onset of Atrazine causes depletion of nonprotein Meydani and

methylmercury neurotoxicity sulfhydryls Hathcock 1984

Vitamin C deficiency Increased severity of Antioxidant properties of Vitamin C and Yamini and Sleight

and methylmercury neurological damage protection against oxidative damage caused by 1984

mercury

Vitamin E and Increased survival and Protection is possibly related to antioxidant Welsh 1979

methylmercury decreased toxicity properties of Vitamin E affording protection

against oxidative damage caused by mercury

Potassium dichromate Synergistic inhibition of Mercuric chloride and potassium dichromate are Baggett and Berndt

and mercuric chloride renal transport both toxic to renal proximal tubule 1984
6-1

6. HAZARD IDENTIFICATION AND DOSE-RESPONSE ASSESSMENT

6.1 Background

Risk assessments done by U.S. EPA follow the paradigm established by the National Academy

of Sciences (NRC 1983). This entails a series of interconnected steps including hazard identification,

dose response assessment, exposure assessment and risk characterization. Two processes, hazard

identification and dose response are the focus of this chapter. Volume IV of this Report presents the

assessment of exposure to mercury emissions in the atmosphere, and Volume VII covers the risk

characterization.

Hazard identification poses the following questions: is the agent in question likely to pose a

hazard to human health; and what types of adverse effects could be expected as a consequence of the

exposure to the agent. Dose-response assessment uses available human, experimental animal and in vitro

data to estimate the exposure level or dose which is expected to produce and adverse health effect. In

accomplishing the aims of risk assessment U.S. EPA applies published Guidelines for Risk Assessment.

6.1.1 Hazard Identification

U.S. EPA has published Guidelines for hazard identification in three areas: developmental

effects, germ cell mutagenicity, and carcinogenic effects. Guidelines for assessment of reproductive

effects were finalized while this Report to Congress was in the process of review; these have not been

applied to the Mercury Study. The specific categorizations for each of those endpoints described in

published guidelines are discussed below. For general, systemic noncancer effects, there is no structured

process resulting in a categorization; instead, the hazard identification step is included in the

dose-response assessment process, wherein a critical effect is selected.

6.1.1.1 Developmental Effects

Guidelines for hazard identification in the area of developmental effects were developed by U.S.

EPA in 1986 and subsequently revised (U.S. EPA 1989, 1991). The Guidelines direct that data from all

available relevant studies be considered, whether the studies indicate a potential hazard or not. Preferred

data are from human studies, when available, and animal studies. The revised guidelines do not use an

alphanumeric scheme such as that given in the carcinogenicity guidelines. Instead two broad categories

are used to characterize the health-related data base: Sufficient Evidence and Insufficient Evidence. The

Guidelines define Sufficient Human Evidence as follows:

"...data from epidemiologic studies (e.g., case control and cohort) that provide

convincing evidence for the scientific community to judge that a causal relationship is or

is not supported. A case series in conjunction with strong supporting evidence may also

be used".

Sufficient Experimental Animal Evidence/Limited Human Data is described in the following way:

"The minimum evidence necessary to judge that a potential hazard exists generally

would be data demonstrating an adverse developmental effect in a single, appropriate,

well-conducted study in a single experimental animal species. The minimum evidence
6-2

needed to judge that a potential hazard does not exist would include data from

appropriate, well-conducted laboratory animal studies (at least two) which evaluated a

variety of the potential manifestations of developmental toxicity, and showed no

developmental effects at doses that were minimally toxic to the adult."

6.1.1.2 Germ Cell Mutagenicity

The U.S. EPA (1986) has published Guidelines for classification of potential hazard of

mutagenic effects in human germ cells. Evidence from human and animal in vivo and in vitro systems is

considered in the judgement as to which of eight numerical classes of concern most clearly defines the

data on an environmental agent. In general, the hierarchy of preference for data type is the following:

 Data on germ cells are preferred to data on somatic cells;

 In vivo tests are preferred to in vitro;

 Data from tests in eukaryotes are preferred to data from prokaryotes.

The weight-of-evidence categories are these, presented in order of decreasing strength of evidence for

human germ cell mutagenicity.

1. Positive data derived from human germ cell mutagenicity studies.

2. Valid positive results from studies on heritable mutational events (any kind) in

mammalian germ cells.

3. Valid positive results from mammalian germ cell chromosome aberration studies that do

not include an intergeneration test.

4. Sufficient evidence for a chemical's interaction with mammalian germ cells, together

with valid positive mutagenicity test results from two assays systems, at least one of

which is mammalian. The positive results may be both for gene mutations or both for

chromosome aberrations; if one is for gene mutations and the other for chromosome

aberrations, both must be from mammalian systems.

5. Suggestive evidence for a chemical's interaction with mammalian germ cells, together

with valid positive mutagenicity evidence from two assay systems as described under 4.

6. Positive mutagenicity test results of less strength than defined under 4, combined with

suggestive evidence for a chemical's interaction with mammalian germ cells.

7. Non-mutagenic. Although definitive proof of non-mutagenicity is not possible, a

chemical could be classified operationally as a non-mutagen for human germ cells, if it

gives valid negative results for all endpoints of concern.

8. Not classifiable based on inadequate evidence bearing on either mutagenicity or

chemical interaction with mammalian germ cells.
6-3

These categories are intended as guidance in assessing a level of concern for an agent’s

likelihood to be a germ cell mutagen. The three forms of mercury are discussed in Section 6.2.2 in terms

of level of concern rather than an assigned numerical category.

6.1.1.3 Carcinogenic effects

U.S. EPA categorizes the carcinogenic potential of a chemical, based on the overall weight-ofevidence, according to the following scheme.

Group A: Human Carcinogen. Sufficient evidence exists from epidemiology studies

to support a causal association between exposure to the chemical and human cancer.

Group B: Probable Human Carcinogen. There is sufficient evidence of

carcinogenicity in animals with limited (Group B1) or inadequate (Group B2) evidence

in humans.

Group C: Possible Human Carcinogen. There is limited evidence of carcinogenicity

in animals in the absence of human data.

Group D: Not Classified as to Human Carcinogenicity. There is inadequate human

and animal evidence of carcinogenicity or no data are available.

Group E: Evidence of Noncarcinogenicity for Humans. There is no evidence of

carcinogenicity in at least two adequate animal tests in different species or in both

adequate epidemiologic and animal studies.

For specific guidance as to the use of human, animal and supporting data in the above

categorization for cancer, consult the Risk Assessment Guidelines of 1986 (U.S. EPA 1987a).

U.S. EPA has been in the process of revising its Guidelines for cancer risk assessment. The

revised Guidelines will implement the use of narrative categorization. The new guidelines also

encourage greater use of mechanistic data, including information which can be gained from genetic

toxicology. Data which elucidate the mode of action of an agent will also have a direct impact on the

dose response assessment for carcinogenicity. In the past a default procedure for dose response

assessment was most often followed; that of linear low dose extrapolation using an upper bound on the

low dose term of a linearized multistage mathematical model. The revised Guidelines dictate that the

type of low dose extrapolation to be used, if any, be guided by information on the carcinogen's mode of

action. Evidence of genetic toxicity has now become key in making decisions about dose response

assessment.

While the Mercury Study Report to Congress was in preparation, revised Carcinogen Risk

Assessment Guidelines were published in the Federal Register. As the approval process was not final, it

was necessary to apply the existing Guideline's alphanumeric categories; however, an expanded narrative

was done, and the weight of evidence judgement followed closely the revised format for expanded

consideration of mechanistic data.

An application of the proposed revisions to the Carcinogenic Risk Assessment Guidelines was

presented at the 1996 meeting of the Society for Risk Analysis (Schoeny 1996). An abstract of that

presentation is given at the end of section 6.
RfD  (NOAEL or LOAEL )

[Uncertainty Factor (s) x Modifying Factor ]

 mg/kgday

RfC  (NOAELHEC or LOAELHEC )

[Uncertainty Factor (s) x Modifying Factor ]

 mg/m3

6-4

6.1.2 Dose-response Assessment

6.1.2.1 Systemic Noncancer Effects

In the quantification of systemic noncarcinogenic effects, an oral reference dose (RfD), an

inhalation reference concentration (RfC) or both may be calculated. The oral RfD and inhalation RfC are

estimates (with uncertainty spanning perhaps an order magnitude) of a daily exposure to the human

population (including sensitive subgroups) that is likely to be without an appreciable risk of deleterious

health effects during a lifetime. The RfD and RfC are derived from a no-observed-adverse-effect level

(NOAEL), or lowest-observed-adverse-effect level (LOAEL), identified from a subchronic or chronic

study and divided by an uncertainty factor(s) times a modifying factor. The RfD or RfC is calculated as

follows:

The methodologies used to derive the RfD or inhalation RfC require the adjustment of NOAEL

and LOAEL values (whether from experimental animal or human studies) to lifetime exposure

conditions (i.e., 24 hours per day for a lifetime of 70 years). Inhalation RfC methods further require

conversion by dosimetric adjustment or the use of a physiologically-based pharmacokinetic model from

NOAELs and LOAELs observed in laboratory animal experiments to human equivalent concentrations

(HEC). Different default adjustments are made based on whether the observed toxicity is in the upper or

lower respiratory tract or at remote sites, and a NOAEL or LOAEL is derived for use in the (HEC) (HEC)

equation above (U.S. EPA 1990).

Selection of the uncertainty factor (UF) to be employed in the calculation of the RfD/RfC is

based upon professional judgment which considers the entire data base of toxicologic effects for the

chemical. In order to ensure that UFs are selected and applied in a consistent manner, the U.S. EPA

(1994) employs a modification to the guidelines proposed by the National Academy of Sciences (NAS

1977, 1980), as shown in the box on the next page.

As noted in the box, the standard UF for extrapolating from animals to humans has been reduced

to three for the derivation of inhalation RfCs. A factor of three was chosen because, assuming the range

of the UF is distributed log normally, the reduction of a standard 10-fold UF by half (i.e. 10 ) results in 0.5

three. Other UFs can be reduced to three if the situation warrants, based on the scientific judgement of

the U.S. EPA RfD/RfC Work Group (an Agency peer review group). Considerations in the selection of

UFs and/or a modifying factor include, but are not limited to, pharmacokinetics/pharmacodynamics,

concomitant exposures, relevance of the laboratory animal models, species sensitivity, severity of the

effect, potential for recovery, slope and shape of the dose-response curve, exposure uncertainties, quality

of the critical study, and data gaps.
DWEL  (RfD) x (Body weight in kg)

Drinking Water Volume in L/day  mg /L

6-5

Uncertainty Factors Used in RfD/RfC Calculations

Standard Uncertainty Factors (UFs)

Use a 10-fold factor when extrapolating from valid experimental results from studies using prolonged

exposure to average healthy humans. This factor is intended to account for the variation in sensitivity

among the members of the human population. [10 ]H

Use an additional 10-fold factor when extrapolating from valid results of long-term studies on

experimental animals when results of studies of human exposure are not available or are inadequate. This

factor is intended to account for the uncertainty in extrapolating animal data to risks for humans. [10 ] A A

3-fold uncertainty factor is used for extrapolating from inhalation studies on experimental animals to

humans for the derivation of an inhalation RfC. This difference is because dosimetric adjustments reduce

the uncertainty associated with extrapolation between experimental animals and humans.

Use an additional 10-fold factor when extrapolating from less than chronic results on experimental

animals when there are no useful long-term human data. This factor is intended to account for the

uncertainty in extrapolating from less than chronic NOAELs to chronic NOAELs. [10 ]S

Use an additional 10-fold factor when deriving an RfD from a LOAEL instead of a NOAEL. This factor

is intended to account for the uncertainty in extrapolating from LOAELs to NOAELs. [10 ]L

Modifying Factor (MF)

Use professional judgment to determine another uncertainty factor (MF) that is greater than zero and less

than or equal to 10. The magnitude of the MF depends upon the professional assessment of scientific

uncertainties of the study and data base not explicitly treated above, e.g., the completeness of the overall

data base and the number of species tested. The default value for the MF is 1.

From the RfD, a Drinking Water Equivalent Level (DWEL) can be calculated. The DWEL

represents a medium-specific (i.e., drinking water) lifetime exposure at which adverse, noncarcinogenic

health effects are not anticipated to occur. The DWEL assumes 100% exposure from drinking water. 

The DWEL provides the noncarcinogenic health effects basis for establishing a drinking water standard. 

For ingestion data, the DWEL is derived as follows:

where:

Body weight = assumed to be 70 kg for an adult

Drinking water volume = assumed to be 2 L/day for an adult

6.1.2.2 Developmental Effects

For agents considered to have sufficient evidence for developmental toxicity it is appropriate to

consider calculation of a quantitative dose-response estimate. In general, a threshold is assumed for the

dose-response curve for agents producing developmental toxicity. This is "based on the known capacity
6-6

of the developing organism to compensate for or to repair a certain amount of damage at the cellular

level. In addition, because of the multipotency of cells at certain stages of development, multiple insults

at the molecular or cellular level may be required to produce an effect on the whole organism" (U.S. EPA

1991).

Due to the paucity of human data, dose-response assessment of developmental toxicity is most

often done using animal data. The assessment includes the identification of dose levels associated with

observed developmental effects as well as those doses which apparently produce no adverse effects. The

critical effect is ascertained from the available data. A critical effect is defined as the most sensitive

developmental effect from the most appropriate and/or sensitive mammalian species; LOAEL and

NOAEL determinations are then made. The NOAEL is defined as "the highest dose at which there is no

statistically or biologically significant increase in the frequency of an adverse effect in any of the

possible manifestations of developmental toxicity when compared with the appropriate control group in a

data base characterized as having sufficient evidence for use in risk assessment" (U.S. EPA 1991). The

LOAEL is defined in the following manner: "The LOAEL is the lowest dose at which there is a

statistically or biologically significant increase in the frequency of adverse developmental effects when

compared with the appropriate control group in a data base characterized as having sufficient evidence".

Because of the limitations associated with the use of the NOAEL/LOAEL approach, U.S. EPA is

investigating the use of alternative methods employing more data in a dose-response assessment. One

such approach is the estimation of a benchmark dose (BMD). This approach is based on the use of a

mathematical model to derive an estimate of an incidence level (e.g., 1%, 5%, 10%, etc). This is done by

applying a model to data in the observed range, selecting an incidence level at or near the observed range

(typically 10%), and then determining an upper confidence limit on the modeled curve. The value of the

upper limit, for a 10% incidence, is then used to derive the BMD, which is the lower confidence limit on

dose for that incidence level.

The last step in dose-response assessment is the calculation of a reference dose or reference

concentration for developmental toxicity (RfD or RfC ). This is done by applying appropriate DT DT

uncertainty factors to the LOAEL, NOAEL, or BMD. Uncertainty factors generally include the

following:

 10 for interspecies variation (animal to human)

 10 for intraspecies variation

Additional factors may be applied to account for other areas of uncertainty, such as identification

of a LOAEL in the absence of a NOAEL. In this case, the factor may be as much as 10 fold, depending

on the sensitivity of the endpoints evaluated in the data base. An uncertainty factor is generally not used

to account for duration of exposure when calculating the RfD . If developmental toxicity is the critical DT

effect for the chronic RfD, an additional uncertainty factor (for study duration) may be used. Modifying

factors may be used to deal with the degree of confidence in the data base for the agent being evaluated. 

For a discussion of application of uncertainty factors to BMDs, see U.S. EPA (1991).

6.1.2.3 Germ Cell Mutagenicity

According to U.S. EPA (1986), a dose-response assessment of an agent's potential for human

germ cell mutagenicity can presently be done using only data from in vivo heritable germ cell tests. This

will remain the case until such time as other assays are demonstrated to have an equivalent predictability

for human effects. The usable tests are, thus, limited to the following: morphological specific locus and
6-7

biochemical specific locus assays; and heritable translocation tests. Data from such assays are generated

from exposures much higher than those expected for humans as a consequence of environmental

exposure. Estimation of extent of human risk is done by extrapolating the observed mutation frequency

or phenotypic effects downward to the expected human exposure range. Available data and mechanistic

considerations are used in the choice of the dose-response model and extrapolation procedure.

6.1.2.4 Carcinogenic Effects

Mathematical models can be used, if data are sufficient, to calculate the estimated excess cancer

risk associated with either the ingestion or inhalation of the contaminant if toxicologic evidence leads to

the classification of the contaminant as one of the following: A, Known Human Carcinogen; B, Probable

Human Carcinogen; or C, Possible Human Carcinogen. The data used in these estimates usually come

from lifetime exposure studies using animals. In order to estimate the potential cancer risk for humans

from animal data, animal doses must be converted to equivalent human doses. This conversion includes

correction for noncontinuous exposure for less-than-lifetime exposure studies and for differences in size. 

The factor to compensate for the size difference should be determined from appropriate experimental

data. In the absence of such data, a default value should be used, such as the cube root of the ratio of the

animal and human body weights. A default assumption is that the average adult human body weight is

70 kg, that the average water consumption of an adult human is 2 L of water per day, and that the

average adult breathes 20 m of air per day. 3

For contaminants with a carcinogenic potential, chemical levels are correlated with a

carcinogenic risk estimate by employing a cancer potency (unit risk) value together with the assumption

for lifetime exposure. The cancer unit risk has generally been derived by assuming low dose linearity

and applying a mathematical model such as a linearized multistage model with a 95% upper confidence

limit. Cancer risk estimates have also been calculated using other models such as the one-hit, Weibull,

logit and probit. There is little basis in the current understanding of the biologic mechanisms involved in

cancer to suggest that any one of these models is able to predict risk more accurately than any other. 

Because each model is based upon differing assumptions, the estimates derived for each model can differ

by several orders of magnitude.

The scientific data base used to calculate and support the setting of cancer risk rate levels has an

inherent uncertainty that is due to the systematic and random errors in scientific measurement. In most

cases, only studies using experimental animals have been performed. Thus, there is uncertainty when the

data are extrapolated to humans. When developing cancer risk rate levels, several other areas of

uncertainty exist, such as the incomplete knowledge concerning the health effects of contaminants in

environmental media, the impact of the experimental animal's age, sex and species, the nature of the

target organ system(s) examined and the actual rate of exposure of the internal targets in experimental

animals or humans. Dose-response data usually are available only for high levels of exposure and not for

the lower levels of exposure closer to where a standard may be set. When there is exposure to more than

one contaminant, additional uncertainty results from a lack of information about possible interactive

effects.

6.2 Hazard Identification for Mercury

Because there are no U.S. EPA guidelines for hazard identification of systemic noncancer

effects, this section does not include a discussion of systemic noncancer effects.
6-8

6.2.1 Developmental Effects

6.2.1.1 Elemental Mercury

Data for developmental effects of elemental mercury are detailed in Section 3.1.3.11 (Tables 3-

25, 3-26 and 3-27). Human studies are inconclusive. The study by Mishinova et al. (1980) provided

insufficient experimental detail to permit evaluation of an exposure-response relationship. Sikorski et al.

(1987) found an increase in reproductive failure among 57 dental professionals by comparison to

controls. This reproductive failure (described as spontaneous abortions, stillbirths or congenital

malformations) was significantly correlated with exposure level. Maternal toxic signs were not reported. 

The study was limited by the small population and the lack of description of the control group. These

findings were not reproduced in Ericson and Kallen's 1989 study of 8157 infants born to dental

professionals in Sweden. When compared to the general population, there was no increase in

malformations, abortions or stillbirths. Exposure data were limited in this study.

There are four animal studies evaluating potential developmental effects associated with

exposure to elemental mercury. In Baranski and Szymczyk (1973), female rats (strain not specified)

were exposed to 2.5 mg/m mercury vapor for 6 to 8 weeks before fertilization or for 3 weeks prior to 3

mating and on days 720 of gestation. In the first experiment, mortality among pups was increased in

the exposed group, and there were changes in pup organ weights (decreased kidney and liver weight and

increased ovary weight). In the second exposed group, mean number of live pups was decreased;

mortality among pups was 100% by day 6 post partum. There were signs of frank toxicity in the dams

including spasms, tremors and death. Information is taken from an English translation of this Polish

paper.

Steffek et al. (1987) is reported in abstract. Rats (strain not specified) were exposed to 0.1, 0.5

or 1.0 mg/m mercury for either the entire gestation period or for days 1015. No effects on resorption 3

or gross abnormality were seen in the low-dose group. Exposure to the mid and high doses for days

1015 resulted in increased numbers of resorption (5/41 and 7/71, respectively; denominators are

presumed to be numbers of litters -- not specified in text); exposure for the entire gestational period

resulted in gross defects in 2/115 fetuses in the low dose and increased resorption (19/38) in the high

dose. Maternal and fetal weight was decreased in the group exposed to 1.0 mg/m for the entire gestation 3

period. No statistical analyses were reported in the abstract.

Two studies in rats focused on behavioral changes consequent to inhalation of elemental mercury

during development. In the first, Danielsson et al. (1993) exposed pregnant Sprague-Dawley rats to

mercury vapor at 1.8 mg/m for either 1 or 3 hours on gestation days 1114 and 1720. There were no 3

signs of toxicity in the dams and offspring of treated animals were no different from controls on the

following measures: body weight; clinical signs; pinna unfolding; surface righting reflex development;

tooth eruption; and results of a negative geotaxis test at days 7, 8 or 9 post partum. Male rats exposed in

utero were significantly hypoactive by comparison to controls at 3 months and hyperactive at 14 months. 

Exposed males were impaired in a test of habituation to novel environments and showed decreased

ability to learn a maze. They were not different from controls in a circular swim test administered at 15

months of age. Females were tested only in the spontaneous motor activity tests; treated females were no

different from controls on this measure.

These results were similar to those reported by Fredriksson et al. (1992). In this instance rats

were exposed postnatally on days 1117 of age to 0.05 mg Hg/m for either 1 or 3 hours/day. High-dose 3

rats showed increased activity (rearing) at 2 months but had decreased activity by comparison to controls
6-9

at 4 months. Low-dose rats were no different from controls at 2 months; at 4 months this group showed

increased total activity and decreased rearing. In the spatial learning test administered at 6 months lowdose rats showed increases in time to complete the task. High-dose animals were observed to have

increases in both time to complete the task and in numbers of errors. Data were not reported on gender

differences in behavior as a result of exposure to mercury vapor.

Both of these studies involved exposure during critical developmental periods, one pre-natal and

one post-natal prior to sexual maturity. Both showed differences from controls (by ANOVA) on one of

four major manifestations of developmental effects listed in the Guidelines for Developmental Toxicity

Risk Assessment (U.S. EPA 1991); namely, functional deficits, in this case in locomotion and learning.

In the Danielsson et al. (1993) paper, these deficits were observed in male offspring in the absence of

maternal toxicity, which according to the Guidelines raises the level of concern. The studies suggest that

the observed effects are not reversible. Latency to reach a platform in the circular swim maze was

significant in the high-dose group at 15 months but not at 7 months, and total activity was decreased in

the low-dose group and increased in the high-dose group at 14 months.

The Guidelines specify that for a judgement of Sufficient Experimental Animal

Evidence/Limited Human Data the minimum data set is the following:

" The minimum evidence necessary to judge that a potential hazard exists generally

would be data demonstrating an adverse developmental effect in a single, appropriate,

well conducted study in a single experimental animal species."

As the data set for elemental mercury consists of two appropriate studies albeit with minimal

group sizes and two incompletely reported studies suggestive of effect, the judgement of Sufficient

Experimental Animal Evidence/Limited Human Data is the most appropriate.

6.2.1.2 Inorganic Mercury

Data on developmental effects of mercuric chloride are found in Section 3.2.3.8 (Tables 3-46 and

3-47). There is one study in mice of developmental effects of inhaled mercuric chloride and none in

humans. Selypes et al. (1984) reported increases in delayed ossification and dead or resorbed fetuses as

a consequence of exposure of CFLP/N mice to 0.17 and 1.6 mg Hg/m as mercuric chloride in an aerosol 3

for 4 hours on days 912 of gestation. There were no statistical analyses, reporting of blood mercury

levels or evaluation of maternal toxicity.

Developmental effects following oral exposure to methylmercury are reported in five oral studies

in rats and hamsters (Table 3-47). McAnulty et al. (1982) is reported in abstract. Oral (not further

specified) mercuric chloride was administered on days 615 of gestation at doses of 6, 9, 12, or 18 mg

Hg/kg-day. This resulted in decreased fetal and placental weights in fetuses in the 6 mg/kg-day group

and malformation at the highest dose. The authors concluded that inorganic mercury was a

developmental toxicant only at doses which were maternally toxic.

Rizzo and Furst (1972) treated Long Evans rats with 0 or 2 mg Hg/kg-day as mercuric oxide on

gestation days 5, 12 or 19. Effects noted were growth retardation and inhibition of eye formation in the

group treated on day 5. No statistical analyses were reported, nor were blood mercury levels given.

Pritchard et al. (1982a) reported in abstract results of treating rats (strain not specified) with

mercuric chloride at 3.0, 6.0 or 12.0 mg Hg/kg-day for about 32 days (gestation day 15 until 25 days
6-10

postpartum). Effects included decreased pup weight and weight gain with a LOAEL of 6.0. In another

experiment reported in an abstract, Pritchard et al. (1982b) exposed female rats to mercuric chloride

doses of up to 18 mg Hg/kg-day before mating and during gestation. High implantation loss was

observed with exposure to 9 mg Hg/kg-day and higher. Embryonic and fetal development was reported

to be unaffected with doses up to 9 mg Hg/kg-day. The abstracts presented insufficient details, and there

was no reporting of statistical analyses.

Gale (1974) gavaged female hamsters (10/group) on gestation day 8 with mercuric acetate at the

following doses: 2.5, 5.0, 16.0, 22.0, 32.0, 47.0, or 63.0 mg Hg/kg-day. There were 3 control animals. 

A variety of malformations and growth effects were noted in animals treated with 16 mg/kg-day or

higher. The authors also treated hamsters via other routes. Their evaluation of efficacy in production of

fetal effects was i.p. > i.v. > s.c. > oral. Maternal toxicity included weight loss, diarrhea, slight tremor,

somnolence, tubular necrosis and hepatocellular necrosis (dose levels not specified). There was

insufficient detail reported for determination of a NOAEL for dams.

In addition to studies of oral or inhalation administration of inorganic mercury there are several

studies which indicate that mercury salts cause developmental toxicity when delivered i.p., s.c. or i.v.

routes (Bernard et al. 1992; Gale and Ferm 1971; Gale 1974, 1981; Kajiwara and Inouye 1986, 1992;

Kavlock et al. 1993). In Gale (1981), wherein exposure was of six strains of hamster to mercuric acetate,

s.c., there was no description of maternal toxicity. In Kavlock et al. (1993) (s.c., rats, mercuric acetate)

fetal effects were noted at doses above the lowest observed maternally toxic dose. Kajiwara and Inouye

(1986) reported their opinion that in mice injected i.v. with mercuric chloride, fetal toxicity was related

to maternal toxicity. In their 1992 study, there was no determination whether implantation loss in

mercuric chloride exposed dams was due to fetal toxicity or to maternal uterine dysfunction. The effects

reported by Bernard et al. (1992) can be better characterized as a transitory nephrotic effect rather than a

developmental deficit.

Each of these studies is limited in its usefulness for assessment of the risk of inorganic mercury

to cause human developmental toxicity. The data base as a whole suggests an effect of inorganic

mercury at doses as low as 2 mg Hg/kg-day. The data, however, are considered insufficient for risk

assessment based on any single study or on the database as a whole (Insufficient Evidence, in the

language of the Guidelines).

6.2.1.3 Methylmercury

Data for developmental effects of methylmercury are presented in Section 3.3.3.8 (Tables 3-66,

3-67 and 3-68); studies are primarily by the oral route and none by the inhalation route. Human studies

of developmental effects include evaluation of children born to mothers exposed to contaminated grain in

Iraq (Amin-Zaki et al. 1976; Marsh et al. 1981, 1987) and contaminated fish in Japan (Harada 1978). 

Effects noted in the Iraqi children included delays in speech and motor development, mental retardation,

reflex abnormalities and seizures. Infants born to mothers ingesting fish from the contaminated

Minamata Bay in Japan appeared normal at birth. Within several months, however, the following effects

were noted: mental retardation, retention of primitive reflexes, cerebellar symptoms, dysarthria,

hyperkinesia, hypersalivation, strabismus and pyramidal symptoms. Histologic examination of brain

tissues of infants from both populations showed a number of signs of pathology. Kjellstrom et al.

(1989), in a study of a population in New Zealand, has observed an inverse correlation between IQ in

children and hair mercury levels in their mothers. In a group of Cree Indians in Quebec, maternal hair

mercury level was correlated with abnormal muscle tone in male children (McKeown-Eyssen et al.

1983).
6-11

Numerous animal studies have demonstrated a variety of developmental effects occurring in rats,

mice and monkeys exposed orally to methylmercury and are presented in Chapter 3 (Table 3-68). 

Developmental effects have been observed in offspring of rats of three strains treated orally with

methylmercury. Developmental effects have also been seen in two strains of mice as well as in guinea

pigs, hamsters and monkeys.

In rodents exposed in utero, decreased fetal weight and increased fetal malformations and deaths

have been reported (Fuyuta et al. 1978, 1979; Inouye and Kajiwara 1988a; Inouye and Murakami 1975;

Khera and Tabacova 1973; Nolen et al. 1972; Reuhl et al. 1981; Yasuda et al. 1985). 

Methylmercury exposure during gestation as well as during the lactation period produces

neurodevelopmental effects (structural and functional alterations) in the exposed pups. Structural effects

include lesions in the brain mantle, corpus callosum, caudate putamen, and cerebellum. In guinea pigs,

early gestational exposures (weeks 35 of pregnancy) resulted primarily in developmental disturbances

of the brain (smaller brains, dilated lateral ventricles, and reduced size of caudate putamen), whereas

later gestational exposures (>week 6 of pregnancy) resulted in widespread neuronal degeneration (Inouye

and Kajiwara 1988b). Functional changes include abnormal tail position during walking, flexion,

hindlimb crossing, decreased locomotor activity, increased passiveness and startle-response, impaired

maze performance, operant behavior, swimming behavior, tactile-kinesthetic function, visual recognition

memory, and temporal discrimination (Bornhausen et al. 1980; Buelke-Sam et al. 1985; Burbacher et al.

1990; Elsner 1991; Geyer et al. 1985; Gunderson et al. 1988; Hughes and Annau 1976; Inouye et al.

1985; Musch et al. 1978; Olson and Boush 1975; Rice 1992; Rice and Gilbert 1990; StoltenburgDidinger and Markwort 1990; Suter and Schon 1986).

While there are limitations to some of these studies (e.g., lack of information on BML, small

study size), the totality of the data base supports a judgment of Sufficient Human and Animal Data for

Developmental Toxicity of methylmercury.

6.2.2 Germ Cell Mutagenicity

6.2.2.1 Elemental Mercury

Data for genotoxicity of elemental mercury are described in Section 3.1.3.13 (Table 3-30). 

Results for an association of somatic cell chromosomal effects with occupational exposure to elemental

mercury are variable. Popescu et al. (1979) and Verschaeve et al. (1976) reported increased incidence of

aberrations or aneuploidy. Most recently Barregard et al. (1991) showed a significant correlation

between cumulative exposure to elemental mercury and micronuclei induction in T-lymphocytes. 

Negative results were reported by Verschaeve et al. (1979) and Mabille et al. (1984). No studies of

mutagenic effect are reported.

Elemental mercury once absorbed is widely distributed throughout the body; there are no data,

however, on elemental mercury in gonadal tissue. Based on both positive and negative findings for

somatic cell chromosomal aberrations in workers, elemental mercury is placed in a group of low

confidence for potential as a human germ cell mutagen.
6-12

6.2.2.2 Inorganic Mercury

Data for genotoxic effects of inorganic mercury are described in Section 3.2.3.10 (Table

3-50). There are no data on inorganic mercury from human germ cell mutagenicity studies or from

studies on heritable mutational events in animals. Anwar and Gabal (1991) reported a statistically

significant increase by comparison to age-matched controls in both chromosomal aberrations and

micronuclei in lymphocytes of workers exposed to mercury fulminate. There was a correlation between

frequency of aberrations and exposure duration. Elemental mercury has been shown to be clastogenic

both in vivo and in vitro. Results of tests for mutagenicity have been variable; generally test results in

prokaryotes are negative for mutagenicity (but may be positive for DNA damage), and results in

eukaryotes are positive. Suter (1975) observed a small, but statistically significant increase in non-viable

implants when female mice were administered mercuric chloride intraperitoneally; the authors were not

certain whether this was a true dominant lethal effect or was attributable to maternal toxicity.

Chromosome aberrations were observed in somatic cells in occupationally exposed humans

(Anwar and Gabal 1991), in somatic cells of mice exposed by gavage (Ghosh et al. 1991), and in Chinese

Hamster Ovary cells treated in vitro (NTP 1993; Howard et al. 1991). Sex-linked recessive mutations

were not observed in Drosophila (NTP 1993), and positive results in a dominant lethal test were

compromised by maternal toxicity (Suter 1975). There are other data for DNA damage and limited data

for gene mutation. Inorganic mercury is less well-distributed in the body than is elemental mercury; it

does not readily pass blood-brain or placental barriers. In one reported study (Jagiello and Lin 1973),

mice treated intraperitoneally were not shown to have an increased incidence of aneuploidy in

spermatogonia. Watanabe et al. (1982), however, showed that while hamsters injected s.c with mercuric

chloride had no increase in aberrations in metaphase II oocytes, there was detectable mercuric chloride in

ovaries and some inhibition of ovulation.

The totality of available data indicates a moderate weight of evidence for germ cell

mutagenicity: sex-linked recessive and dominant lethal results were compromised, but there are positive

results for chromosomal aberrations in multiple systems (including in vivo exposure) and evidence that

mercuric chloride can reach female gonadal tissue.

6.2.2.3 Methylmercury

Summaries of data for genotoxicity of methylmercury are presented in Section 3.3.3.10 (Tables

3-70, 3-71 and 3-72).

Methylmercury appears to be clastogenic but not a potent mutagen. Methylmercury is widely

distributed in the body, breaching both blood-brain and placental barriers in humans. There are data

indicating that methylmercury administered i.p. reaches germ cells and may produce adverse effects. 

Suter (1975) observed a slight reduction in both numbers of implantations and viable embryos in (SEC x

C57Bl)F females which had been mated to treated males. This was not noted in (101 x C3H)F mice. 1 1

When Syrian hamsters were treated intraperitoneally with methylmercury, aneuploidy but not

chromosomal aberrations was seen in oocytes. Sex-linked recessive lethal mutations were increased in

Drosophila melanogaster given dietary methylmercury. Watanabe et al. (1982) noted some decrease in

ovulation in hamsters treated s.c. with methylmercury, further indication that methylmercury is

distributed to female gonadal tissue.

Studies have reported increased incidence of chromosome aberrations (Skerfving et al. 1970,

1974) or SCE (Wulf et al. 1968) in lymphocytes of humans ingesting mercury-contaminated fish or
6-13

meat. Chromosome aberrations have been reported in cats treated in vivo and in cultured human

lymphocytes in vitro. Evidence of DNA damage has been shown in a number of in vitro systems.

As there are data for mammalian germ cell chromosome aberration and limited data from a

heritable mutation study, methylmercury is placed in a group of high concern for potential human germ

cell mutagenicity. All that keeps methylmercury from the highest level of concern is lack of positive

results in a heritable mutation assay.

6.2.3 Carcinogenic Effects

This section presents the critical carcinogenicity studies evaluated by the U.S. EPA for the

weight-of-evidence classification of elemental, inorganic (mercuric chloride) and organic

(methylmercury) forms of mercury. These studies are discussed more completely in Chapter 3 and

summarized in Tables 3-1, 3-31, 3-32, 3-33, 3-52, and 3-53.

6.2.3.1 Elemental Mercury

Human data regarding the carcinogenicity of inhalation of elemental mercury are insufficient to

determine whether such exposures may result in increased cancer incidence. Several studies report

statistically significant increases in lung cancer mortality among groups of exposed workers (Amandus

and Costello 1991; Barregard et al. 1990; Buiatti et al. 1985; Ellingsen et al. 1992). The interpretation of

these studies is limited by small sample sizes, probable exposure to other known lung carcinogens,

failure to consider confounders such as smoking and failure to observe correlations between estimated

exposure and the cancer incidence. A study of dental professionals found a significant increase in the

incidence of glioblastomas (Ahlbom et al. 1986). It is not known whether exposure to mercury, X-rays,

or other potential carcinogens in the workplace contributed to the effects observed. No increase in

cancer mortality was observed among workers exposed to mercury vapor in a nuclear weapons facility

(Cragle et al. 1984), but this study was also limited by the small sample size. No studies were identified

that examined cancer incidence in animals exposed chronically to elemental mercury vapor. These

studies are presented in greater detail in Section 3.1.2.

The overall findings from cytogenetic monitoring studies of workers occupationally exposed to

mercury by inhalation provide very limited evidence of genotoxic effects. Popescu et al. (1979)

compared four men exposed to elemental mercury vapor with an unexposed group and found an

increased number of chromosomal aberrations. Verschaeve et al. (1979) found an increased incidence of

aneuploidy after exposure to low concentrations.

In summary, human epidemiological studies failed to show a correlation between exposure to

elemental mercury vapor and increased cancer incidence, but the studies are limited by confounding

factors. Only one study in animals is reported (Druckrey et al. 1957); tumors were found only at contact

sites, and the study is incompletely reported as to controls and statistics. Animal data are, thus, also

inadequate. Findings from assays for genotoxicity are limited and provide no convincing evidence that

mercury exposure has an effect on the number or structure of chromosomes in human somatic cells. The

most appropriate category is, thus, Group D, not classifiable as to human carcinogenicity.

This classification was reviewed by the Carcinogen Risk Assessment Verification Endeavor

(CRAVE), an Agency Peer Review Work Group. The classification was accepted as appropriate on

March 3, 1994.
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6.2.3.2 Inorganic Mercury

There are no data available on the carcinogenic effects of inorganic mercury (mercuric chloride)

in humans. In animals, there is equivocal evidence of carcinogenicity in rats and mice. In rats gavaged

with mercuric chloride for two years (NTP 1993), survival was significantly reduced in males (17% and

8% survival in low and high-dose males versus 43% survival in controls), indicating that the maximally

tolerated dose (MTD) was exceeded. There was an increased incidence of forestomach squamous cell

papillomas (0/50, 3/50, 12/50 in control, low, and high-dose males, respectively; 0/50, 0/49 and 2/50 in

control, low and high-dose females, respectively). Papillary hyperplasia of the forestomach was also

significantly elevated in both male dose groups and in high-dose females. In addition, the incidence of

thyroid follicular cell carcinomas in treated males (1/50, 2/50 and 6/50 in control, low- and high-dose

males, respectively) showed a significantly positive trend. There were, however, no increases in thyroid

hyperplasia of adenomas; it is not clear that the increase in thyroid carcinomas is a treatment-related

effect. The NTP also considered the forestomach tumors to be of limited relevance to humans; there was

no evidence that these contact site tumors progressed to malignancy.

In a companion study in mice (NTP 1993), there was a significantly increasing trend for renal

tubular cell tumors (adenomas and adenoma carcinomas). No dose groups were statistically significantly

different from the control by pair-wise comparison, although the incidence in the high-dose group was

elevated. There was a significant increase in severe nephropathy in treated animals. The NTP studies

and two nonpositive bioassays are summarized in Section 3.2.2.

In summary, there are no data in humans linking mercuric chloride with carcinogenic effects. 

Data in animals are limited. Focal hyperplasia and squamous cell papillomas of the forestomach as well

as thyroid follicular adenomas and carcinomas were observed in male rats gavaged with mercuric

chloride. In the same study, evidence for increased incidence of squamous cell forestomach papillomas

in female rats and renal adenomas and carcinomas in male mice were considered equivocal. All

increased tumor incidences were observed at what were considered high doses (in excess of the MTD). 

In this context, the relevance of the thyroid tumor to human health evaluation has been questioned; these

tumors are considered to be secondary to the hyperplastic response. Results from in vitro and in vivo

tests for genotoxicity have been mixed with no clear indication of a strong somatic cell genotoxic effect

of mercuric chloride exposure.

Based on the absence of human data and limited data for carcinogenicity in animals, mercuric

chloride is classified as Group C, possible human carcinogen. This classification was reviewed by

CRAVE on March 3, 1994 and found to be appropriate.

6.2.3.3 Methylmercury

The available human data are inconclusive regarding the carcinogenicity of methylmercury in

humans exposed by the oral route. A study of leukemia patients from a rural area in Poland showed a

significantly higher mercury content in hair in the leukemia patients than in healthy unrelated patients or

healthy relatives (Janicki et al. 1987). The population studied was small, and the study did not adjust for

other leukemia risk factors. In addition, two studies of larger populations exposed to methylmercury

during the Minamata incident failed to show increases in leukemia or total cancer incidence (Tamashiro

et al. 1984, 1986). Although one of these studies showed a significant increase in liver cancer incidence,

factors other than mercury exposure were likely contributors to the increase. These epidemiological

studies are presented in greater detail in Section 3.3.2.1.
6-15

Animal studies show some evidence of carcinogenicity in two strains of mice, but studies in rats

have not shown similar results. Male ICR mice given methylmercuric chloride in the diet for up to two

years had significantly increased incidences of renal epithelial adenomas and/or adenocarcinomas

(Hirano et al. 1986; Mitsumori et al. 1981). Similarly, male B6C3F1 mice given methylmercuric

chloride in the diet for up to two years had significantly increased incidences of renal epithelial

carcinomas and adenomas (Mitsumori et al. 1990). In contrast, Sprague-Dawley rats administered

methylmercury in the diet for up to 130 weeks exhibited no increase in tumor incidence (Mitsumori et al.

1983, 1984). Although the dose was lower in the rats than in the mice, a maximally tolerated dose was

achieved in the rat study as evidenced by an approximately 2030% decrease in body weight gain and by

significant increases in renal and neuronal toxicity in both male and female rats at the highest dose

tested. Other studies also failed to show increases in tumor incidence after chronic exposure to

methylmercury (Schroeder and Mitchener 1975; Verschuuren et al. 1976), but these studies were limited

by small sample sizes, failure to achieve a maximally tolerated dose and/or incomplete histopathological

examinations. These studies are presented more completely in Section 3.3.2.2.

In summary, data for carcinogenicity from human studies are considered inadequate. Three

studies that examined the relationship between methylmercury exposure in humans and increased

incidence of cancer were limited by poor study design or incomplete description of methodology or

results. Data from animal studies are considered to provide limited evidence of carcinogenicity. Male

ICR and B6C3F1 mice exposed to methylmercuric chloride in the diet were observed with increased

incidence of renal adenomas, adenocarcinomas and carcinomas. Tumors were observed at a single site,

in a single species and sex. Renal epithelial cell hyperplasia and tumors were observed only in the

presence of profound nephrotoxicity; tumors were suggested to be consequent to reparative changes in

the affected organs. Although genotoxicity test data suggest that methylmercury is clastogenic, there are

also negative tests.

The limited data in animals above support a categorization of Group C, possible human

carcinogen. The CRAVE Work Group accepted this weight-of-evidence judgment as appropriate at its

March 3, 1994 meeting.

6.2.3.4 Application of proposed revision of the Guidelines for Carcinogen Risk Assessment 

Data described in the above three sections were re-evaluated using criteria described in the

proposed revisions to the Guidelines for Carcinogen Risk. Among the changes in the revised guidelines

are emphasis on use of data describing the mode of action of the putative carcinogen, both in weight of

evidence judgements and in decisions as to the most appropriate type of low dose extrapolation. The

revised Guidelines encourage consideration of relevance to human health risk of route and magnitude of

exposure in animal bioassays. Both of these considerations were part of a re-evaluation of the data for

carcinogenicity of the three forms of mercury. Results of a presentation (Schoeny, 1996) on the

application of the revised Guidelines are summarized below.
6-16

Elemental mercury

The likelihood of elemental mercury to be a human carcinogen cannot be determined; data in

humans and animal bioassays are inadequate. Epidemiologic studies, though confounded showed no

correlation between exposure to elemental mercury vapor and carcinogenicity. Animal data were not

positive, but the published study was considered inadequate. The study was done by a route

(intramuscular injection) not relevant to human exposure (inhalation). Genetic toxicity data are limited

and equivocal.

Inorganic (mercuric chloride)

The data for inorganic mercury indicate that it is not likely to be a human carcinogen under

conditions of exposure generally encountered in the environment. There are no data on carcinogenicity in

humans. Findings in animals included squamous cell papillomas of the forestomach and thyroid

follicular cell adenomas and carcinomas in gavaged rats, and renal adenomas and adenocarcinomas in

male mice. All increased tumor incidences were at high doses (in excess of the MTD). Genetic toxicity

data gave a mix of positive and negative response for chromosomal breakage and were equivocal for 

somatic cell point mutations. The mode of action for forestomach tumors appears to be a high dose

effect related to irritation and cytotoxicity. The mode of action for thyroid tumors is not clear; there was

no treatment-related increase in incidence of hyperplasia. There was high mortality in rats from renal

toxicity. Renal toxicity in male mice was less severe; the mode of action of inorganic mercury in

producing renal neoplasms is not clear. Human exposure (other than occupational or accidental

poisoning) is likely to be to low levels of inorganic mercury in water or food plants.

Methylmercury

Methylmercury is not likely to be a human carcinogen under conditions of exposure generally

encountered in the environment. Data in humans were inadequate; interpretation is limited by

inappropriate study design and incomplete descriptions of methodology. Dietary exposure in two strains

of mice resulted in increased renal adenomas and adenocarcinomas. Tumors were observed only in dose

groups experiencing profound nephrotoxicity. Studies in rats exposed to a MTD showed no increased

tumor incidence. Several studies show that methylmercury can cause chromosomal damage in somatic

cells. While evidence is good for chromosomal effects, it does not appear that methylmercury is a point

mutagen. The mode of action in renal tumor induction is likely to be related to reparative changes in the

tissues. Human exposure is likely to be from consumption of contaminated foods especially fish. It is

expected that exposure, even in groups consuming large amounts of fish from contaminated sources, will

be to levels far below those likely to cause the tissue damage associated with tumor formation in

animals.

6.3 Dose-Response Assessment For Mercury

6.3.1 Systemic Noncancer Effects

6.3.1.1 Oral Reference Doses (RfDs)

Elemental mercury

Metallic mercury is only slowly absorbed by the gastrointestinal tract (~0.01%) and because of

this is thought to be of no toxicological consequence (Klaassen et al. 1986) when ingested. Further

discussion of an RfD for this form of mercury is not presented.
6-17

Inorganic mercury (mercuric chloride)

An RfD for inorganic mercury of 3x10 mg/kg-day has been verified by the RfD/RfC Work -4

Group. The critical effect serving as the basis for the RfD is kidney toxicity due to an auto-immune

disease caused by the accumulation of IgG antibodies in the glomerular region of the kidneys. 

On October 26 and 27 of 1987, a panel of mercury experts met at a Peer Review Workshop

convened by U.S. EPA for the purpose of reviewing outstanding issues concerning the health effects and

risk assessment of inorganic mercury (U.S. EPA 1987). The panel participants are listed in Appendix C. 

Five consensus conclusions and recommendations were agreed to as a result of this workshop; these are

presented in Table 6-1. The RfD was determined using data on autoimmune glomerulonephritis

observed in rats. Based on three studies using the Brown-Norway rat, a DWEL value was determined

using studies described below. The Brown-Norway rat is very sensitive to this mercuric mercuryinduced autoimmune effects, although this effect has also been demonstrated in other strains of rats and

other species of experimental animals (Andres 1984; Bernaudin et al. 1981; Hultman and Enestrom

1992). The Brown-Norway rat is believed to be a good surrogate for the study of mercury-induced

kidney damage in sensitive humans (U.S. EPA 1987b). The glomerulonephritis is characterized by

deposition of anti-glomerular basement membrane antibodies (IgG) in renal glomeruli and after

prolonged exposure is often accompanied by proteinuria and, in some cases, nephrosis (Druet et al.

1978).

LOAEL values were identified from three individual studies. In Druet et al. (1978), BrownNorway rats were exposed to mercuric chloride via subcutaneous injection, 3 times/week, for 8 weeks. 

The dose levels administered were 0, 0.1, 0.25, 0.5, 1.0 and 2.0 mg Hg/kg, and there were

620 animals/group. An additional group of animals received 0.05 mg Hg/kg for 12 weeks. (The

number of animals/sex was not stated.) Druet and colleagues measured antibody formation (using a

fluoresceinated sheep anti-rat IgG antiserum) and urinary protein levels. Proteinuria occurred at doses

0.1 mg/kg (LOAEL); the proteinuria was considered a highly deleterious effect, as it frequently led to

hypoalbuminemia and even death. A LOAEL for lifetime exposure was calculated to be 0.226 mg/kgday, using the following conversion:

0.05 mg/kg x 3 days/7 days x 0.739 [HgCl 2  Hg ] x 100% absorption/7% 2+

= 0.226 mg Hg/kg-day

In a 60-day study conducted by Bernaudin et al. (1981), Brown-Norway rats (5/group) were

force-fed 0 or 3 mg/kg/week mercuric chloride. At the end of the 60 days, there were no classic

histological abnormalities in the kidneys of treated animals. Using immunofluorescence, however, IgG

deposition was evident in all of the treated rats, and weak proteinuria was noted in 3/5 dosed animals. A

lifetime LOAEL was calculated to be 0.317 mg Hg/kg-day. Dose conversion was done in the following

manner:

3 mg/kg x 1 day/7 days x 0.739 [HgCl  Hg ] 2

2+

= 0.317 mg Hg/kg-day
RfD  DWEL x 2 L/day

70 kg bw

6-18

Table 6-1

Consensus Decisions of Peer Review Panel

 The most sensitive adverse effect for mercury risk assessment is formation of mercuric

mercury-induced autoimmune glomerulonephritis. The production and deposition of IgG

antibodies to the glomerular basement membrane can be considered the first step in the

formation of this mercuric mercury-induced autoimmune glomerulonephritis.

 The Brown-Norway rat should be used for mercury risk assessment. The Brown-Norway rat

is a good test species for the study of Hg -induced autoimmune glomerulonephritis. The 2+

Brown-Norway rat is not unique in this regard (i.e., this effect has also been observed in

rabbits). 

 The Brown-Norway rat is a good surrogate for the study of mercury-induced kidney damage

in sensitive humans. For this reason, the uncertainty factor (for interspecies variability) used

to calculate criteria and health advisories (based on risk assessments using the Brown Norway

rat) should be reduced by 10-fold.

 Hg absorption values of 7% from the oral route and 100% from the subcutaneous route 2+

should be used to calculate criteria and health advisories.

 A DWEL of 0.010 mg/L was recommended based on the weight of evidence from the studies

using Brown Norway rats and limited tissue data.

Similar results were obtained by Andres (1984). Five Brown-Norway rats were exposed to

3 mg/kg mercuric chloride via gavage 2 times/week for 60 days. In this same study, Lewis rats (n=2)

were also exposed using the same dosing regimen. After 60 days, the kidneys of all treated animals

appeared normal histologically, and no proteinuria was reported in any treated animals; IgG deposition in

the renal glomeruli was demonstrated using immunofluorescence in Brown-Norway rats. No antibody

deposition was noted in the Lewis rats. The lifetime LOAEL was determined to be 0.633 mg Hg/kg-day. 

Dose conversion was done in the following manner:

3 mg/kg x 2 days/7 days x 0.739 [HgCl  Hg ] 2

2+

= 0.633 mg Hg/kg-day

As the result of intensive review of these and other studies, as well as the discussions of the

panel of mercury experts convened for this purpose, a recommended DWEL of 0.01 mg/L was derived

from the LOAELs above, and, subsequently, the oral RfD value was back-calculated:

RfD = 0.010 mg/L x 2L/day/70 kg bw

= 0.0003 mg/kg bw/day
6-19

The RfD for inorganic mercury was reviewed by the RfD/RfC Work Group which reached

consensus for verification on November 16, 1988. The Work Group agreed to application of an

uncertainty factor of 1000 to the LOAELs above (which ranged from 0.23 to 0.63 mg Hg/kg-day). The

uncertainty factor was composed of a 10-fold each for subchronic to chronic and LOAEL to NOAEL

extrapolation, and an additional 10-fold factor for both animal to human and sensitive populations. The

resulting RfD of 3x10 mg/kg-day was given high confidence based on the weight of the evidence from -4

the studies using Brown Norway rats and the entirety of the data base.

A literature search for the years 1988 to 1994 has been conducted and recently reviewed

(September 1994). The NTP (1993) study was among those considered. A rat NOAEL of 0.23 mg

Hg/kg administered dose has been identified for renal effects for the 6-month portion of the study. A

description of the NTP gavage study has been included in the summary information for IRIS. U.S. EPA

concluded that no change in the RfD for inorganic mercury is needed at this time.

Methylmercury

U.S. EPA has on two occasions published RfDs for methylmercury which have represented the

Agency consensus for that time. These are described in the sections below. The original RfD of 0.3

μg/kg/day was determined in 1985. The current RfD of 0.1 μg/kg/day was established as Agency

consensus in 1995. At the time of the generation of the Mercury Study Report to Congress, it became

apparent that considerable new data on the health effects of methylmercury in humans were emerging. 

Among these are large studies of fish or fish and marine mammal consuming populations in the

Seychelles and Faroes Islands. Smaller scale studies are in progress which describe effects in

populations around the U.S. Great Lakes. In addition, there are new evaluations, including novel

statistical approaches and application of physiologically-based pharmacokinetic (PBPK) models, to

published work described in section 3.3.1.1 of this volume.

 As much of this new data has either not yet been published or have not yet been subject to

rigorous review, it was decided that it was premature for U.S. EPA to make a change in the 1995

methylmercury RfD at this time. This decision was approved by the Science Advisory Board (SAB), a

public advisory group providing extramural scientific information and advice to the Administrator and

other officials of the Environmental Protection Agency. The SAB is structured to provide balanced,

expert assessment of scientific matters relating to problems facing the Agency. Their report makes the

following statement. 

In general, from the standpoint of looking at human health effects and the uncertainties, the

draft report is a very good document and an important step forward in terms of bringing the relevant

information together into one place for the first time. The current RfD, based on the Iraqi and New

Zealand data, should be retained at least until the on-going Faeroe and Seychelles Islands studies have

progressed much further and been subjected to the same scrutiny as has the Iraqi data.

The SAB report continues:

Investigators conducting two new major prospective longitudinal studies--one in the Seychelles

Islands the other in the Faeroe Islands--have recently begun to publish findings in the literature and are

expected to continue releasing their findings during the next 2-3 years. These studies have advantages

over those cited in the previous paragraph in that they have much larger samples sizes, a larger number

of developmental endpoints, potentially more sensitive developmental endpoints, and control a more

extensive set of potential confounding influences. On the other hand, the studies have some limitations

in terms of low exposures (to PCBs in the Faeroes) and ethnically homogenous societies. Since only a
6-20

small portion of these new data sets have been published to date and because questions have been raised

about the sensitivity and appropriateness of the several statistical procedures used in the analyses, the

Subcommittee concluded that it would be premature to include any data from these studies in this report

until they are subjected to appropriate peer review. Because these data are so much more

comprehensive and relevant to contemporary regulatory issues than the data heretofore available,

once there has been adequate opportunity for peer review and debate within the scientific

community, the RfD may need to be reassessed in terms of the most sensitive endpoints from these

new studies.

An inter-agency process, with external involvement will be undertaken for the purpose of

reviewing these new data, their evaluations, and the evaluations of existing data. An outcome of this

process will be an assessment by U.S. EPA of its RfD for methylmercury to determine if a change is

warranted.

Former RfD

A hazard identification and dose-response assessment was proposed for methylmercury in 1980

(U.S. EPA 1980) and later verified by the RfD/RfC Work Group on December 2, 1985. This assessment

was subsequently included on U.S. EPA’s Integrated Risk Information System (IRIS). The critical

effects were multiple central nervous system (CNS) effects including ataxia and paresthesia in

populations of humans exposed to methylmercury through consumption of contaminated grain

(summarized by Clarkson et al. 1975, Nordberg and Strangert 1976 and WHO 1976); see study

descriptions in Section 3.3.

The RfD for methylmercury was determined to be 3x10 mg/kg-day, based on a LOAEL of -4

0.003 mg/kg-day (corresponding to 200 μg/L blood concentration) and an uncertainty factor of 10 used

to adjust the LOAEL to what is expected to be a NOAEL. An additional uncertainty factor of 10 for

sensitive individuals for chronic exposure was not deemed necessary at the time of the RfD's

verification, as the adverse effects were seen in what was regarded as a sensitive group of individuals,

namely adults who consumed methylmercury-contaminated grain.

Medium confidence was ascribed to the choice of study, data base and RfD. The blood levels

associated with the LOAEL were well supported by more recent data, but neither the chosen studies nor

supporting data base described a NOAEL. Medium confidence indicates that new data may change the

assessment of the RfD.

Since the time of verification, several submissions to IRIS have questioned the value of this RfD,

and, specifically, whether or not this RfD is protective against developmental effects. Subsequent to the

RfD verification, the effects in Iraqi children of in utero exposure to methylmercury were reported by

Marsh et al. (1987). Discussion of the methylmercury RfD by the RfD/RfC Work Group was reported in

1992 and 1994. Consensus for verification of the RfD described below was reached in January of 1995.
d g/day  CxbxV

Axf

6-21

Current U.S. EPA RfD

Determination of critical effect

Marsh et al. (1987) was chosen as the most appropriate study for determination of an RfD

protective of a putative sensitive subpopulation; namely infants born to mothers exposed to

methylmercury during gestation. This paper describes neurologic abnormalities observed in progeny of

women who consumed bread prepared from methylmercury-treated seed grain while pregnant (See

Chapter 3 for study description). Among the signs noted in the infants exposed during fetal development

were cerebral palsy, altered muscle tone and deep tendon reflexes as well as delayed developmental

milestones (i.e., walking by 18 months and talking by 24 months). Each child in the study was examined

by two neurologists who scored observed effects on a scale for severity ranging from 0 to 11. The data

collected by Marsh et al. (1987) summarize clinical neurologic signs of 81 mother and child pairs. From

x-ray fluorescent spectrometric analysis of selected regions of maternal scalp hair, concentrations

ranging from 1 to 674 ppm mercury were determined, then correlated with clinical signs observed in the

affected members of the mother-child pairs. Among the exposed population there were affected and

unaffected individuals throughout the exposure range.

Method employed for determination of critical dose

In order to quantitate an average daily mercury ingestion rate for the mothers, hair

concentrations were determined for periods during gestation when actual methylmercury exposure had

occurred. This procedure is possible since hair grows an average rate of 1 cm/month (Al-Shahristani et

al. 1976) and since Iraqi women wear their hair very long; appropriate samples were, thus, available for

the period of gestation when exposure occurred.

A number of laboratory studies support a correlation between hair concentrations and concurrent

blood concentrations. Some variation in the ratio exists; a ratio of 250:1 (μg mercury/mg in hair:μg

mercury/ml of blood) was used to derive the RfD critical dose. A more complete discussion for the

choice of this ratio is provided in the next section.

The hair concentration at a hypothetical NOAEL for developmental effects was determined by

application of a benchmark dose approach (see subsequent section for discussion of methods and data

used). The analysis used the combined incidence of all neurological effects in children exposed in utero

as reported in the Marsh et al. (1987) study. A Weibull model for extra risk was used to determine the

benchmark dose of 11 ppm mercury in maternal hair (11 mg/kg hair). This was converted to 44 g/L

blood using the above 250:1 ratio.

11 mg/kg hair / 250 = 44 g/L blood

To obtain a daily dietary intake value of methylmercury corresponding to a specific blood concentration,

factors of absorption rate, elimination rate constant, total blood volume and percentage of total mercury

that is present in circulating blood were taken into account. Calculation was by use of the following

equation based on the assumptions that steady state conditions exist and that first-order kinetics for

mercury are being followed.
d  cxbxV

A x f x bw

d  44 g/L x 0.014 days1

x 5L

0.95 x 0.05 x 60 kg

d  1.1 g/kgday

6-22

Where:

d = daily dietary intake (expressed as ug of methylmercury)

C = concentration in blood (expressed as 44 ug/liter)

b = elimination constant (expressed as 0.014 days )-1

V = volume of blood in the body (expressed as 5 liters)

A = absorption factor (expressed as a unitless decimal fraction of 0.95)

f = fraction of daily intake taken up by blood (unitless, 0.05)

Solving for d gives the daily dietary intake of mercury which results in a blood mercury concentration of

44 g/L. To convert this to daily ingested dose (g/kg-day) a body weight of 60 kg was assumed and

included in the equation denominator.

The dose d (1.1 g/kg-day) is the total daily quantity of methylmercury that is ingested by a 60 kg

individual to maintain a blood concentration of 44 g/L or a hair concentration of 11 ppm.

The rationales for use of specific values for equation parameters follow.

Hair to blood concentration ratio. The hair:blood concentration ratio for total mercury is

frequently cited as 250:1 expressed as g mercury/g hair to g mercury/ml of blood. Ratios reported in

the literature range from 140 to 416, a difference of about a factor of 3. Table 6-2 provides the results of

12 recent studies in which hair to blood ratios were calculated for a variety of human populations. 

Differences in the location of hair sampled (head versus chest and distance from scalp) may contribute to

the differences observed. Variability in the hair-blood relationship for mercury concentration can also be

attributed to the fact that unsegmented hair analysis gives a time-weighted average of mercury exposure,

while analysis of mercury in blood reflects a much shorter period average of exposure. As much as a 3-

fold seasonal variation in mercury levels was observed in average hair levels for a group of individuals

with moderate to high fish consumption rates, with yearly highs occurring in the fall and early winter

(Phelps et al. 1980; Suzuki et al. 1992). The relatively high ratio reported by Tsubaki (Table 6-2) may

have reflected the fact that mercury levels were declining at the time of sampling so that the hair levels

reflect earlier, higher blood levels. Cernichiari et al. (1995a) reported a maternal hair:blood ratio of

416:1 for residents of the Seychelles Islands. The authors remarked that while this ratio was high,

statistical uncertainties do not permit a judgement as to whether it is truly outside the range reported in

WHO (1990) recapitulated in Table 6-2. Phelps (1980) obtained multiple blood samples and sequentially

analyzed lengths of hair from individuals. Both hair and blood samples were taken for 339 individuals in

Northwestern Ontario. After reviewing the various reports for converting hair concentrations to blood

concentrations, the analysis in the Phelps (1980) paper was selected by the Agency RfD/RfC Work

Group because of the large sample size and the attention to sampling and analysis that was made. The
6-23

ratio Phelps observed between the total mercury concentration in hair taken close to the scalp and

simultaneous blood sampling for this group was 296:1. To estimate the actual ratio the authors assumed

that blood and hair samples were taken following complete cessation of methylmercury intake. They

also assumed a half-life of methylmercury in blood of 52 days and a lag of 4 weeks for appearance of the

relevant level in hair at the scalp. Phelps also determined that 94% of the mercury in hair is

methylmercury. Based on these assumptions, they calculated that if the actual hair:blood ratio were

200:1, they would have observed a ratio of 290. Based on these and other considerations, Phelps states

that the actual ratio is "probably higher than 200, but less than the observed value of 296." As the

authors point out, one-third of the study population was sampled during the rising phase of seasonal

variation ( and two-thirds or more in the falling phase). Phelps et al. (1980) had assumed that all were

sampled in the falling phase. This fact would tend to result in a lower observed ratio; therefore, the

actual average is likely to be greater than 200. It was concluded by U.S. EPA that a midpoint value of

250 is acceptable for the purpose of estimating average blood levels in the Iraqi population.

Fraction of mercury in diet that is absorbed (A). After administration of radiolabeled

methylmercuric nitrate in water to 3 healthy volunteers, uptake was reported to be >95%. (Aberg et al.

1969). This value is supported by experiments in human volunteers conducted by Miettinen et al.

(1971). These researchers incubated fish liver homogenate with radiolabeled methylmercury nitrate to

produce methylmercury proteinate. The proteinate was then fed to fish for a week; the fish were killed,

cooked and fed to volunteers after confirmation of methylmercury concentration. Mean uptake exceeded

94%. Based upon these experimental results an absorption factor of 0.95 was used in these calculations.

Fraction of the absorbed dose that is found in the blood (f). There are three reports of the

fraction of absorbed methylmercury dose distributed to blood volume in humans. Kershaw et al., (1980)

reported an average fraction of 0.059 of absorbed dose in total blood volume, based on a study of 5 adult

male subjects who ingested methylmercury-contaminated tuna. In a group of 9 male and 6 female

volunteers who had received Hg-methylmercury in fish, approximately 10% of the total mercury body 203

burden was present in one liter of blood in the first few days after exposure; this dropped to

approximately 5% over the first 100 days (Miettinen et al. 1971) In another study, an average value of

1.14% for the percentage of absorbed dose in one kg of blood was derived from data on subjects who

consumed a known amount of methylmercury in fish over a 3-month period (Sherlock et al. 1984). 

Average daily intake in the study ranged from 43 to 233 μg/day, and there was a dose-related effect on

percentage of absorbed dose that ranged from 1.03% to 1.26% in one liter of blood. Each of these values

was multiplied by 5 to yield the total amount in the blood compartment, as there are approximately 5

liters of blood in an adult human body (0.01 x 5 = 0.05). The value 0.05 has
6-24

Table 6-2

Available Data on Hair:Blood Ratio (total Hg)

Reference Ratio Subjects (g/L) Scalp Hair (ppm) (mm)

Hair to Number Hg Range in Distance

Blood of Whole Blood to Hg Range in Length

Hair Samples

Sumari et al., 1969 140 50 5-270 1-57 -- -- 1

Soria et al., 1992 218 16 2.4-9.1 0.15-20 -- at scalp

Tejning, 1967 230 51 4-110 1-30 -- axillary 1

Skerfving, 1974 230 60 44-550 1-142 5 at scalp

Haxton et al., 1979 250 173 0.4-26 0.1-11.3 20 --

Tsubaki, 1971b 260 45 2-800 20-325 -- -- 2

Birke et al., 1972 280 12 4-650 1-180 5 at scalp 2 3

Den Tonkelaar et al., 1974 280 47 1-40.5 <0.5-13.2 -- --

Kershaw et al., 1980 292 5 -- -- 5 at scalp 4

Phelps et al., 1980 296 339 1-60 1-150 10 at scalp

Sherlock et al., 1982 367 98 1.1-42.3 0.2-21 24 --

Cernichiari et al., 1995 416 740 0.5-26.7 10 at scalp

Tsubaki, 1971a 370 ~25 -- -- longer tuft -- 1 1

 As cited in Berglund et al. 1971 1

 As cited in WHO, 1976 2

 Ratio of methylmercury in hair to methylmercury in blood 3

 Based on repeated measurements at different time points (3-8 ratios per individual) of the ratio of 5 mm hair segments to 4

 corresponding 2-week average blood levels (assuming hair growth of 1.1 cm/month).

-- = Not reported

been used for this parameter in the past by other groups; e.g.., Berglund et al. (1971) and WHO (1990). 

A value of 0.05 was used for "f" in the above equation.

Elimination constant (b). Several studies reported clearance half-times for methylmercury from

blood or hair in the range of 35-189 days (Miettinen 1972; Kershaw et al. 1980; Al-Shahristani et al.

1974; Sherlock et al. 1984). Two of these studies included the Iraqi population exposed during the

19711972 incident. A value reported in Cox et al. (1989) was derived from the study group which

included the mothers of the infants upon which this risk assessment is based. The average elimination

constant of the 4 studies is 0.014; the average of individual values reported for 20 volunteers ingesting

from 42 to 233 μg mercury/day in fish for 3 months (Sherlock et al. 1982) is also 0.014. A value of

0.014 days was, thus, used for term "b" in the above equation. -1

Volume of blood in the body (V). That blood volume is 7% of body weight has been determined

by various experimental methods. There is an increase of 20% to 30% (to about 8.5 to 9%) during
6-25

pregnancy (Best 1961). Specific data for the body weight of Iraqi women were not found. Assuming an

average body weight of 58 kg and a blood volume of 9% during pregnancy, a blood volume of 5.22 liters

was derived. In the equation on page 6-19, term "V" was taken to be 5 liters.

Body weight. While the critical endpoint for the RfD is developmental effects in offspring, the

critical dose is calculated using parameters specific to the mothers who ingested the mercury

contaminated grain. Data on body weights of the subjects were not available. A default value of 60 kg

(rounded from 58) for an adult female was used.

Grouping of data

Data used in the U.S. EPA benchmark dose calculation were excerpted from the publication,

Seafood Safety (NAS 1991). The tables of incidence of various clinical effects in children that were

provided in this document readily lent themselves to the benchmark dose modeling approach. The

continuous data for the Iraqi population that were reported in Marsh et al. (1987) were placed in five

dose groups, and incidence rates were provided for delayed onset of walking, delayed onset of talking,

mental symptoms, seizures, neurological scores above 3, and neurological scores above 4 for affected

children. Neurologic scores were determined by clinical evaluation for cranial nerve signs, speech,

involuntary movement, limb tone strength, deep tendon reflexes, plantar responses, coordination,

dexterity, primitive reflexes, sensation, posture, and ability to sit, stand and run. Table 6-3 shows the

input data for the modeling procedure for effects found in children. Incidence data for each of the

adverse effects in children were taken directly from Table 6-11, Seafood Safety (NAS, 1991). The

effects of late walking, late talking, and neurologic scores greater than 3 were also combined for

additional analysis. Table 6-4 shows the incidence data for each of the effects observed in adults as

grouped in Table 6-13 of the Seafood Safety document.

Adjustments for background incidence

As an adjustment for background rates of effects, the benchmark dose estimates for

methylmercury were calculated to estimate the dose associated with "extra risk." Another choice would

have been to calculate based on "additional risk." Additional risk (AR) is defined as the added incidence

of observing an effect above the background rate relative to the entire population of interest, AR = [P(d)-

P(0)]/1. In the additional risk calculation, the background rate is subtracted off, but still applied to the

entire population, including those exhibiting the background effect, thus in a sense "double counting" for

background effects. It can be seen that extra risk (ER) is always mathematically greater than or equal to

additional risk, ER = [P(d)-P(0)]/[1-P(0)], and is thus a more conservative measure of risk (whenever the

background rate is not equal to zero). Conceptually, extra risk is the added incidence of observing an

effect above the background rate relative to the proportion of the population of interest that is not

expected 
6-26

Table 6-3

Incidence of Effects in Iraqi Children By Exposure Groupa

Effect 1.37 10 52.53 163.38 436.60

Late Walking 0 2 2 3 12

Late Talking 2 1 3 4 11

Mental Symptoms 1 0 1 3 4

Seizures 0 0 1 2 4

Neurological Scores >3 3 1 4 3 9

Neurological Scores >4 0 1 2 2 6

All Endpoints 4 3 6 8 14

 N 27 14 13 12 15

 From Table 6-11 of Seafood Safety; Dose is geometric mean in ppm maternal hair. a

Table 6-4

Incidence of Effects in Iraqi Adults by Exposure Groupa

Effect 50 350 750 1500 2500 3500 4500

Paresthesia 2 1 8 10 20 14 7

Ataxia 1 0 2 8 15 17 7

Visual Changes 0 0 4 9 14 10 6

Dysarthria 1 1 1 4 6 13 6

Hearing Defects 0 0 1 0 3 65

Deaths 0 0 0 0 0 32

 N 21 19 19 17 25 17 7

 From Table 6-13 of Seafood Safety; Dose is geometric mean in ppb blood. a

to exhibit such an effect. Extra risk is then more easily interpreted than additional risk, because it applies

the additional risk only to the proportion of the population that is not represented by the background rate. 
P(d)  A0  (1A0)(1exp[A1  d A2],

6-27

Extra risk has been traditionally used in U.S. EPA's cancer risk assessments (Anderson et al. 1983) and is

discussed in detail in a report on the benchmark dose by U.S. EPA's Risk Assessment Forum (U.S. EPA 1995).

Derivation of a benchmark dose

Benchmark dose estimates were made by calculating the 95% lower confidence limits on doses

corresponding to the 1%, 5% and 10% extra risk levels using a quantal Weibull model (K.S.Crump

Division of ICF Kaiser International). The Weibull model was chosen for the benchmark dose

calculations for the methylmercury data as recent research suggests it may be the best model for

developmental toxicity data (Faustman et al., 1994). The form of the quantal Weibull that was used is

the following:

where d = dose, 

A0 = background rate = 0.12468, 

A1 = slope = 9.47 × 10 , and -3

A2 = shape parameter = 1.000. 

For each endpoint and for the combined endpoints, the incidence of response was regressed on

the dose. A Chi-squared test of goodness-of-fit was used to test the null hypothesis (Ho) that the

predicted incidence was equal to the observed incidence, so that Ho would be rejected for p-values less

than 0.05.

Results for individual effects and all effects combined for children exposed in utero are given in

Table 6-5; results for adults are given for comparison in Table 6-6. For calculation of the lower bound

on the 10% risk level, A0 = 0.12468, A1 = 9.470230 x 10 , A2 = 1.00000. The RfD/RfC Work Group -3

chose the benchmark (lower bound on the dose for 10% risk) based on modeling of all effects in

children. Recent research (Allen et al. 1994a, b) suggests that the 10% level for the benchmark dose

roughly correlates with a NOAEL for developmental toxicity data. Note that this conclusion was based

on controlled animal studies and on calculation of additional risk. Both the polynomial and Weibull

models place a lower 95% confidence limit on the dose corresponding to a 10% risk level at 11 ppm hair

concentration for methylmercury. The benchmark dose rounded to 11 ppm was used in the calculation of

the RfD.

Dose groupings other than those used in Seafood Safety were also done and benchmark doses run

as above for comparison. Both density-based grouping and uniform concentration intervals were used.

The local density of observations relative to the mercury level in hair was analyzed using a

density estimation algorithm (smooth function in S-PLUS for Windows, Ver. 3.1; S-PLUS Guide to

Statistical and Mathematical Analysis). The function estimates a probability density for the distribution

of a variable by calculating a locally-weighted density of the observations. That is, the function

estimates the probability that an observation will be near a specific value based on how the actual values

are clustered. In this case, the function was used to estimate the probability density for an observation in

the neighborhood of any given maternal hair mercury concentration. The density plot is shown in Figure

6-1. The peaks represent relatively greater numbers of data points than the troughs in the vicinity of the

associated hair mercury concentrations.
6-28

Table 6-5

Methylmercury Benchmark Dose Estimates (ppm hair)

Maximum Likelihood Estimates and 95% Lower Confidence Limits from Weibull Model

Incidence of Effects in Children (Marsh et al. 1987)

Effect G-O-F

0.01 0.05 0.10 a

MLE 95% CL MLE 95% CL MLE 95% CL P-Value

Late Walking 3.3 2.1 16.7 10.9 34.3 22.4 0.16

Late Talking 4.7 2.4 22.1 12.3 43.8 25.3 0.79

Mental Symptoms 12.0 6.4 61.0 32.8 125.4 67.5 0.63

Seizures 11.8 6.7 60.4 34.3 124.2 70.5 0.86

Neuro Score >3 5.6 3.3 28.8 17.0 59.1 34.9 0.58

Neuro Score >4 8.1 4.6 41.4 23.7 84.9 48.7 0.48

All Endpoints 1.6 1.1 8.3 5.4 17.1 11.1 0.94

 Goodness-of-fit p-value for testing the null hypothesis, Ho: Predicted Incidence = Observed Incidence. a

Table 6-6

Methylmercury Benchmark Dose Estimates (ppb blood)

Maximum Likelihood Estimates and 95% Lower Confidence Limits from Weibull Model

Incidence of Effects in Adults (Bakir et al. 1987)

Effect G-O-F

0.01 0.05 0.10 a

MLE 95% CL MLE 95% CL MLE 95% CL P-Value

Paresthesia 45.3 14.3 169.0 73.2 302.2 150.5 0.36

Ataxia 330.4 140.8 652.9 369.7 882.0 564.7 0.22

Visual Changes 64.1 25.4 249.1 129.9 453.6 266.9 0.26

Dysarthria 728.9 235.9 1265.4 621.8 1614.4 949.8 0.41

Hearing Defects 1462.9 535.2 2137.5 1202.8 2527.1 1705.6 0.53

Deaths 2226.3 1106.8 3007.2 2167.3 3434.2 2797.0 0.83

 Goodness-of-fit p-value for testing the null hypothesis, Ho: Predicted Incidence = Observed Incidence. a
ppm Hg in hair

density

1 5 10 50 100 500

0.10 0.15 0.20

6-29

Figure 6-1

Density of Data Points Relative to Mercury

Concentration in Hair for Iraqi Cohort Data

The density distribution is characterized by four distinct peaks. Exposure dose groups were

defined as trough-to-trough intervals with the peak values taken as the nominal value for each interval. 

The nominal dose-group value, concentration ranges, and incidence of combined developmental effects

are given in Table 6-7. A benchmark dose was calculated from the incidence of all effects as grouped in

Table 6-7. The lower 95% confidence interval on the benchmark dose for the 10% response is 13 ppm,

compared to the 11 ppm value used as the basis for the RfD.

Another alternative dose grouping approach was to divide the entire exposure range into four

equal log-dose intervals. The geometric midpoint of each interval was taken as the nominal value for the

interval. The nominal dose-group value, concentration ranges, and incidence of combined

developmental effects are given in Table 6-8. The benchmark calculated as the lower bound on the 10%

incidence for all effects is 10.3 ppm, compared to the 11 ppm used for the RfD.

Table 6-7

Density-Based Dose Groupings

Nominal Dose (ppm) Dose Range (ppm) Incidence

1.18 1 - 4 5/27

10.6 5 - 28 3/16

78.8 29 - 156 10/17

381 157 - 674 18/21
6-30

Table 6-8

Uniform Dose Groupings

Nominal Dose (ppm) Dose Range (ppm) Incidence

2.25 1 - 5 5/28

11.5 6 - 25 3/14

58.6 26 - 132 9/17

299 133 - 674 19/22

Two other analyses were done. In the first, data on males and females were grouped as

published in Seafood Safety, and the Weibull model was as for the data in Table 6-9. The lower 95%

confidence interval on the benchmark dose for the 10% response for males only was 10ppm and for

females only, 11ppm. The last analysis consisted of fitting all data (for males and females ) on all

endpoints in Table 6-9 without grouping . When the model was restricted such that the Weibull power

would drop below 1, the lower 95% bound on the dose for the 10% response was 11ppm. For an

unrestricted model the benchmark dose was. Table 6-9 lists all equivalent benchmarks (lower 95%

bounds on a 10% effect level for all measured endpoints) calculated by U.S.EPA on the data of Marsh et

al. (1987).

Table 6-9

Benchmark Dosed Calculated on Data from Marsh et al. (1987)

Data Grouping Benchmark Dose (ppm maternal hair)

Grouping in Seafood Safety 11

Density-based grouping 13

Uniform concentration intervals 10

Males only (Seafood Safety groups) 10

Females only (Seafood Safety groups) 10

Individual data points (restricted model) 11

Uncertainty and modifying factors

A composite uncertainty factor of 10 was used. This uncertainty factor was applied for

variability in the human population, in particular the wide variation in biological half-life of

methylmercury and the variation that occurs in the hair to blood ratio for mercury. In addition, the factor

accounts for lack of a two-generation reproductive study and lack of data for possible chronic

manifestations of the adult effects (e.g., paresthesia that was observed during gestation). The default

value of one was used for the modifying factor.
RfD  Benchmark Dose

UF x MF

 1.1 ug/kgday

10

 1 x 104mg/kgday

6-31

Calculation of the oral RfD for methylmercury

In this instance the RfD was calculated using the following equation:

Confidence in the oral RfD for methylmercury

The principal study (Marsh et al. 1987) is a detailed report of human exposures with quantitation

of methylmercury by analysis of specimens from affected mother-child pairs. A strength of this study is

that the quantitative data are from the affected population and quantitation is based upon biological

specimens obtained from affected individuals. A threshold was not easily defined; extended application

of modeling techniques were needed to define the lower end of the dose-response curve. This may

indicate high variability of response to methylmercury in the human mother-child pairs or

misclassification of assigning pairs to the cohort. Confidence in the supporting data base and confidence

in the RfD were considered medium by the RfD/RfC Work Group.

An analysis of uncertainty in an RfD based on the Iraqi data is found in Appendix D of this

volume. Discussions of areas of uncertainty can also be found in Volume VII, Risk Characterization.

Choice of Benchmark or NOAEL as the basis for the RfD

Estimates of threshold levels for neurotoxicity have been performed by WHO (1990) using data

from the Niigata episode and the Iraqi poisoning. In the exposures in Japan, hair levels associated with

thresholds for neurotoxicity were estimated by WHO to be approximately 100 ppm. Estimates of

threshold levels associated with paresthesia in the Iraqi episode indicate that the threshold level for

paresthesia is approximately 25 to 40 mg (total body burden). This corresponds to blood levels of

approximately 250 to 400 g/L and hair levels of approximately 50 g/g. Thresholds (total body

burden) estimated by Bakir et al. (1973) for other neurotoxic signs were 55 mg for ataxia, 90 mg for

dysarthria (difficulty with speech), 170 mg for deafness, and 200 mg for death.

A number of additional studies of human populations generally support the dose range of the

benchmark dose level for perinatal effects. The designs for these studies as well as summaries of results

are given in section 3.3.1.1. A few of the studies have data suitable for calculation of benchmark doses

as was done for the results of Marsh et al. (1987). Table 6-10 compares one published benchmark dose

(on the New Zealand population) as well as several others calculated for the Mercury Study Report to

Congress. Table 6-12 provides a compilation of NOAELs and LOAELs determined from inspection of

human data sets as well as other published comparable measures. 

A recent analysis of the Kjellstrom (Kjellstrom et al. 1986a, b,1989) data was published by

Gearhart et al. (1995). In this analysis the authors used a PBPK model which incorporated a fetal
6-32

compartment. They calculated a benchmark dose on all 28 tests included in the initial study design by

Kjellstrom; this was done assuming values of 1 and 5% for background deficiency in test scores. The

range of benchmark doses calculated was 10 to 31 ppm maternal hair mercury. The authors’ preferred

benchmark was 17 ppm, for an estimated background incidence of 5% and the lower bound on the 10%

risk level (Table 6-10).

Table 6-10

Summary of Benchmark Doses (BMD) Estimated for Methylmercury

Study site Duration of n Endpoint Benchmark Reference

Exposure (ppm maternal

a

hair mercury)

b

Iraq short term 81 all developmental 11 Marsh et al (1987);

(~ 3 mo) effects reported in this document

Marsh et al.(1987)

Iraq short term 81 all effects except 15 Marsh et al (1987);

(~ 3 mo) late walking, late this document

talking

Seychelles long term 789 DDST, abnormal 16 Meyers et al. (1995);

plus questionable this document

scores

New Zealand long term 237 all measures in 28 17 Kjellstrom et al.

tests (1986a, b, 1989);

Gearhart et al. (1995)

a n=number of subjects in analysis

b first reference is to source of data; second is to source of BMD calculation. 

Data have recently been published from the pilot (or cross sectional) study and testing of

children from up to age 29 months in the main or prospective study in the Seychelles (Myers et al.

1995a,b,c,d; Davidson et al. 1995). The range of maternal hair mercury levels for the pilot study was 0.6

to 36.4 ppm; the range for the main study was 0.5 to 26.7 ppm. The Among tests administered to the 60

month old children in the cross-sectional study was the Denver Developmental Screen Test (DDST)

(Myers et al. 1995b) which was included largely to provide a point of comparison with other population

studies (e.g. Kjellstrom et al. 1986a, b, 1989). The frequency of abnormal plus questionable scores was

analyzed by multiple logistic regression which showed an association for increased frequency of nonnormal scores with increasing maternal hair mercury. Results for the abnormal plus questionable scores

(Table 6-11) were used to calculate a benchmark dose as a lower 95% limit on a 10% effect level. The

resulting estimate is 16 ppm maternal hair. The study authors have commented that the main study,

which was done under more controlled circumstances, did not show any relationship between DDST-R

results and maternal hair mercury.
6-33

Table 6-11

Results of Revised Denver Developmental Screening Test (DDST) 

Administered to Seychellois Children in Cross-sectional Studya

Result Maternal Hair Mercury, ppm

0-3 >3-6 >6-9 >9-12 >12 total

normal 114 209 169 107 122 721

(92.7) (92.5) (91.4) (93.0) (87.1) (91.4)

abnormal 0 1 1 0 13

questionable 9 16 15 8 17 65

abnormal plus 9 17 16 8 18 68

questionable (7.3) (7.5) (8.7) (7.0) (12.9) (8/6)

a Data from Myers et al. (1995b)

In a recent publication, Gearhart et al. (1995) proposed a RfD in the range of 0.8 to 2.5 μg/kgday based on their analysis of effects in a population of children in New Zealand. This population was

assumed to be exposed in utero to methylmercury as a consequence of high fish consumption by their

mothers. Gearhart et al. (1995) estimated that a maternal intake of methylmercury in the range of 0.8 to

2.5 μg/kg-day corresponded to a NOAEL for developmental effects. These results support the U.S. EPA

estimate of maternal intake of 1 μg/kg-day methylmercury corresponding to a benchmark dose of 11 ppm

mercury in hair for developmental effects, prior to applying an uncertainty factor. The primary area of

disagreement between Gearhart et al. (1995) and the U.S. EPA is in the use of an uncertainty factor. The

authors felt that no uncertainty or modifying factors were needed as the NOAEL was calculated on

effects in a sensitive subpopulation. U.S. EPA applied a 10-fold uncertainty factor to account for

interindividual variation in the human population (particularly in hair to blood mercury ratio) and for

lack of certain types of data.

The NOAEL was implicitly defined by Gearhart et al. (1995) as the lower 95% confidence

interval on the dose associated with a 10% change in the test scores from the New Zealand study

(Kjellstrom et al. 1989). That is, the benchmark dose calculation was performed on continuous variables

by contrast to the binary variables from the Iraqi study (Marsh et al. 1987) used by the U.S. EPA. The

qualitative equivalence of the two kinds of benchmark doses has not been established; both presumably

represent a minimum risk level as does a NOAEL. 

Using a hockey stick parametric dose response analysis of the data on delayed walking in the

Iraqi children, Cox et al. (1989) concluded that the "best statistical estimate" of the threshold for health

effects was 7.3 ppm mercury in hair with a 95% range of uncertainty between 0 and 14. A more recent

analysis of the same data (Cox et al. 1995) focuses on the importance of four data points termed

influential for the estimation of the population threshold. The newer analysis indicates that when a

background response rate of 4% is assumed, the threshold estimate is 9 ppm maternal hair (Table 6-12). 

The authors indicate that dose-response analyses based on the late walking endpoint are unreliable

because it relies on four influential observations in the data set from Marsh et al. (1978). The data points

in question are the only responders below 150 ppm (mercury in hair). In particular, Cox et al. (1995)

state that the four observations are isolated from the remainder of the responders and would be expected

to have considerable influence on threshold estimate. This conclusion is based on a visual interpretation

of a plot of the data (Figure 2 in Cox et al. 1995). Based on visual inspection of the same figure, an

argument could
6-34

Table 6-12

Estimates of No Observed Adverse Effect Levels (NOAELs) and

Lowest Observed Adverse Effects Levels (LOAELs) from Human Studies

Study Site Exposure n Endpoint Estimate Estimate Reference

type in ppm a

mercury

in hair

b

Iraq in utero 84 All developmental NOAEL 13 Marsh et al.

and post (1981); Marsh et

partum al. (1981)

/short term

Iraq in utero 81 All developmental NOAEL 7-10 Marsh et al.

/short term (1987); this

document

Iraq in utero 81 delayed walking LOAEL 14 Marsh et al.

/short term (1987); this

document

Iraq in utero 81 delayed walking best 7.3 Marsh et al.

/short term estimate (1987); Cox et al.

of (1989).

threshold

Iraq in utero 81 delayed walking best 9 Marsh et al.

/short term estimate (1987); Cox et al.

of (1995)

threshold

New in utero 237 IQ tests (WISC-R, threshold 5-15 Kjellstrom et al.

Zealand /long term TOLD) (1986a, b, 1989);

Lipfert (1994)

Canada in utero 247 deep tendon reflex LOAEL 2-15 McKeowin -

(Cree /long term Eyssen (1983);

population) Lipfert (1994).

Amazon adult/long 29 visual LOAEL 20 Lebel et al. (1995);

term discrimination this document

Niigata adult 430 Minimata disease threshold 25 Kinjo et al. (1995);

women/ this document

long term
Table 6-12 (continued)

Estimates of No Observed Adverse Effect Levels (NOAELs) and

Lowest Observed Adverse Effects Levels (LOAELs) from Human Studies

Study Site Exposure n Endpoint Estimate Estimate Reference

type in ppm a

mercury

in hair

b

6-35

Iraq, Cree, in utero multiple effects LOAEL 10-20 Marsh et al.

New /short and (1981), McKeowin

Zealand long term -Eyssen (1983),

Kjellstrom et al

(1986a, b, 1989);

Hoover et al 1997.

Seychelles in utero 217 Preschool LOAEL 12 - 36 Myers et al.

(pilot, 66 /long term Language Scale (1995a); this

month old (auditory document

children) comprehension)

Seychelles in utero 736 BSID NOAEL 12- 26.7 Davidson et al.

(main, 29 /long term 1995; this

month old document

children)

Seychelles in utero 736 Activity level, LOAEL 12- 26.7 Davidson et al.

(main, 29 /long term boys 1995; this

month old document

children)

Faroes in utero 917 Neuropsychologica LOAEL <10 Grandjean et al.

/long term l tests (1997); Grandjean

et al. (1997)

a Threshold and best estimate defined by author of estimate.

b First reference is to source of data; second is to source of analysis. 

be made that the separation is not that marked considering the first eight responders. No quantitative

sensitivity analysis was performed to investigate the effect of removing one or more of these data points. 

Cox et al. (1995) point out that if the four points are assumed to represent background, then the threshold

for late walking would be greater than 100 ppm. It would seem unlikely, however, that these

observations represent background given that no responses were observed in the 37 individuals with

lower levels of exposure. It should be noted that the U.S. EPA benchmark dose was done on incidence

of all effects, rather than on late walking only.

Crump et al. (1995) reanalyzed data from the Iraqi methylmercury poisoning episode presented

in Marsh et al. (1987). In their analysis, Crump et al. (1995) reported that the statistical upper limit of

the threshold could be as high as 255 ppm. Furthermore, the maximum likelihood estimate of the

threshold using a different parametric model was presented as virtually zero. These and other analyses

demonstrated that threshold estimates based on parametric models exhibit high statistical variability and

model dependency, and are highly sensitive to the precise definition of an abnormal response.
6-36

Using a statistical analysis for trend that does not require grouping of the data, Crump et al.

(1995) demonstrated that the association between health effects and methylmercury concentrations in

hair is statistically significant at mercury concentrations in excess of about 80 ppm. In addition, Crump

et al. calculated benchmark doses by applying dose-response models to each of the three endpoints: late

walking, late talking and neurological score. Unlike the benchmark calculations made by the U.S. EPA

(1994), these analyses did not involve grouping of the data into discrete dose groups, nor did they require

dichotomizing continuous responses like age first walked into "late walking" or "no walking." Their

calculation of the 95% lower bounds on the hair concentration corresponding to an additional risk of

10% ranged from 54 ppm to 274 ppm mercury in hair. Crump et al. (1995) concluded that the trend

analyses and benchmark analyses provided a sounder basis for determining RfDs than the type of hockey

stick analysis presented by Cox et al. (1989). They felt that the acute nature of the exposures, as well as

other difficulties with the Iraqi data, present limitations in the use of these data for a chronic RfD for

methylmercury.

The Cox et al. (1995) and Crump et al. (1995) analyses deal primarily with one endpoint, late

walking. This appears to be the most sensitive of the endpoints described in Marsh et al. (1978). Both

Cox et al. and Crump et al., as well as the U.S. EPA analysis in Appendix D of Volume V, show

considerable uncertainty in thresholds estimated from the data on late walking. The peculiar nature of the

uncertainty, in this case, makes it difficult to distinguish between 7 ppm maternal hair mercury and 114

ppm as a best (maximum likelihood) estimate for the threshold. Cox et al. (1995) attribute this bimodal

uncertainty to four influential observations between 14 ppm and 60 ppm isolated from the remainder of

the responders beginning at 154 ppm; they do not present arguments, other than visual, for censoring

these data. Crump et al. (1995) show that changing the definition of late walking from greater than 18

months to 18 months or greater eliminates the bimodal uncertainty with a best threshold estimate of 230

ppm. The implication in both analyses is that the background incidence of late walking, as reported in

other studies, is not consistent with the lower thresholds. While this is true, the use of historical controls

for this analysis may not be appropriate, given the relatively large number of observations at low

exposure levels in the Iraqi cohort; 33% of the observations were at hair mercury concentrations

considered to be background levels (3 ppm or less).

Late walking, as assessed in the exposed Iraqi population (Marsh et al. 1978), is almost certainly

a valid indicator of methylmercury toxicity but may be unreliable as the sole basis for detailed doseresponse analysis. The primary reason for this may be the uncertainty in maternal recall for both birth

date and date of first walking. The uncertainty in this particular case could be quite large, given the lack

of recorded information. The primary impact of this kind of uncertainty would be on the response

classification of individuals at the upper bound of normal (18 months for first walking) and at the lower

bound of abnormal. The lowest abnormal first walking time presented in Marsh et al. (1978) was 20

months. The impact of assuming uncertainty in the classification of the observations in these two groups

is large given the large number of observations in the two groups (19 data points at 18 months and 8 data

points at 20 months). The analysis in Appendix D to Volume V of the Mercury Study Report to

Congress shows that thresholds estimated for late walking are unstable when classification uncertainty is

considered. The same kind of subjective uncertainty is applicable to the late talking endpoint, as well. 

The thresholds for late talking, however, are much more stable, statistically, as there are fewer

observations that are near the normal/abnormal threshold value of 24 months.

McKeown-Eyssen et al. (1983) observed a positive association between abnormal tendon

reflexes in boys and increasing maternal hair mercury. This was a study of 234 Cree children between

the ages of 12 to 30 months residing in northern Quebec communities. Average maternal hair mercury

for boys and girls was 6 ppm; the maximum was 24 ppm and 6% of the population had hair mercury

levels in excess of 20ppm. A LOAEL of between 2-15 ppm maternal hair mercury was calculated by

Lipfert et al. (1996). These authors felt that the McKeown-Eyssen et al. (1983) study provides only

marginal support for a LOAEL above 10 ppm; they concluded that part of the significance of the adverse

effects rests on the in (in their opinion) inappropriate inclusion of two observations of increased muscle
6-37

tone in the 2-15 ppm group. In a later paper, Hoover et al. (1997) reported that a LOAEL range of 10-20

ppm from this study was warranted.

Marsh et al. (1995) have published results of a study conducted between 1981 and 1984 in

residents of coastal communities of Peru. The prospective study was of 131 child-mother pairs; testing

for potential effects of fetal methylmercury exposure was patterned after the study of children exposed in

utero in Iraq. Peak maternal hair methylmercury ranged between 1.2 to 30 ppm with a geometric mean

of 8.3 ppm. These authors showed no effects of methylmercury on measures similar to those performed

on the Iraqi children (including time of first walking and talking). A NOAEL (in the absence of a

LOAEL) from this study would be 30 ppm maternal hair mercury (Table 6-12). 

 

Lebel et al. (1996) published a study of 29 young adults (less than 35 years of age) who resided

in villages on the Tapajos River, about 200 Km downstream of Amazonian gold-mining sites. Hair

mercury ranged from 5.6 ppm to 38.4 ppm; methymercury constituted between 72.2% and 93.3% of the

total. The authors found that decreases in several measures of visual acuity (color discrimination loss,

contrast sensitivity and visual field reduction) were related to increased hair mercury. The authors note

that constriction of the visual field has been reported in other instances of mercury intoxication. 

Inspection of the data presented as charts indicates a LOAEL of about 20ppm.

Kinjo et al.(1995) have published an analysis of the relationship between hair mercury

concentration and the incidence of Minamata disease in Niigata, Japan. Hair samples were collected in

1965 and analyzed by a colorimetric procedure. The population used in the statistical analyses consisted

of 147 males and 430 females. The authors felt that the colorimetric method used to determine hair

mercury was not reliable at levels below 20 ppm. Data on individuals with hair mercury less than 20

ppm were excluded from the analysis; these data, however were included in Figure 3 of the paper. 

Minamata disease was defined by criteria established for receipt of compensation by the Japanese

government medical committee (Tamashiro et al 1985). Both hockey stick and logit models were

applied to obtain a range of threshold values of 24.7 to 49.3 for females; inspection of the graphed data

in Figure 3 indicates a threshold for Minamata disease in females between 20 and 30 ppm. For males the

threshold estimates are between 43 and 48 ppm.

A pilot study (essentially a cross-sectional study) of developmental effects in a seafood

consuming population in the Seychelles Islands focused on all children born between February 1989 and

February 1990. A total of 804 mother-infant pairs were observed to have maternal hair mercury in the

range of 0.6 to 36.4 ppm with a median of 6.6 ppm (Myers et al. 1995b). Children were tested once

between the ages of 5 and 109 weeks of age. An association was found for in utero mercury exposure

and DDST-R abnormal plus questionable scores combined (Myers et al 1995b). A benchmark dose was

derived on this data set as had been done on the Kjellstrom et al. (1986a, b, 1989) data. The BMD (95%

lower bound on the 10% effect level) is 16 ppm maternal hair mercury (Table 6-10).

A subset of the pilot cohort (217 children) was evaluated at 66 months using the McCarthy

Scales of Children’s Abilities, the Preschool Language Scale, and tests from the Woodcock-Johnson

Tests of Achievement that were appropriate to the children’s age (Myers et al. 1995a). The median

maternal hair mercury for this group was 7.11 ppm. Mercury exposure (measured as maternal hair

mercury) was negatively associated with four endpoints: the McCarthy General Cognitive Index and

Perceptual Performance subscale; and the Preschool Language Scale Total Language and Auditory

Comprehension subscale. When statistically determined outliers and points considered to be influential

were removed from the analyses, statistical significance of the association remained only for auditory

comprehension. If a decrease in the scores on the auditory comprehension test is considered an adverse

effect, a LOAEL from this study would be in the range of 12-36 ppm maternal hair mercury.

The main study was designed to be prospective; children were evaluated at 6.5, 19, 29 and 66

months of age (data on the 66 month old children have not yet been published) (Shamlaye et al, 1995) . 
6-38

In the group evaluated at 6.5 months, median maternal hair mercury was 5.9 ppm with a range of 0.5 ppm

to 26.7 ppm. No association with maternal hair mercury was found for any endpoint tested. (Myers et al.

1995c). Evaluations at 19 and 26 months were done on groups of 738 and 736 individuals, respectively

(Davidson et al. 1995). Median maternal hair mercury was 5.9 ppm, and the range was 0.5 to 26.7 ppm.

Children were evaluated with the BSID at 19 months of age. No effects of mercury exposure were seen

on outcome of tests administered at 19 months. At 29 months, children were administered the BSID as

well as the Bayley Infant Behavior Record. At 29 months there was an association between mercury

exposure and decreased activity level in male children only. If the decrease in activity level is considered

an adverse effect, a LOAEL for the study would be in the range of 12-26.7 ppm maternal hair mercury

(Table 6-12). If this study is considered non-positive a NOAEL would be in the range of 12-26.7 ppm

maternal hair mercury. 

The overall conclusion of the studies published to date is that it is yet unclear whether an

association exists between low level mercury exposure and neurologic deficits in children. The study

shows a close correlation between maternal hair mercury and neonatal levels of mercury in brain tissue

(Cernichiari et al. 1995). The authors cautioned in several papers that subtle neurologic and

neurobehavioral effects are more likely to be detected in older rather than younger children. The overall

conclusion of the authors is that their results require careful interpretation, and that an association

between relatively low level mercury exposure in utero and neurologic deficits has not been

demonstrated. 

In 1986 a large study was initiated in the Faroe Islands on neurologic developmental effects of

methylmercury and PCB exposure in utero (Grandjean et al. 1997). Subjects were a group of 917

children born between 1986 and 1987 and examined at about 7 years of age. Mercury was measured in

maternal hair and cord blood, and a subset of cords was evaluated for PCBs. The median maternal hair

mercury concentration was 4.5 ppm, and 13% were greater than 10 ppm (Grandjean et al 1992a). The

geometric mean cord blood mercury concentration for the group of children who completed the

evaluation was 22.8 ppm. Significant negative associations were seen for several neuropsychological

tests. Even after inclusion of covariates with uncertain influence on these tests, multiple regression

analysis indicated that 9/20 measures showed mercury related decrements (p< 0.05, one tailed). 

Application of a Peters-Belson adjustment resulted in significant mercury associations for 11/20

measures. PCB determinations were done on a total of 436 cords, and PCB exposure was included as a

covariate in the regression analyses. This had an effect only on the regression for the Boston Naming

Test. After exclusion of children with maternal hair mercury concentrations above 10 ppm, these

associations remained almost unchanged. The authors concluded that in utero exposure to

methylmercury at levels below 10 ppm maternal hair mercury affects several domains of cerebral 

function; in particular, attention, language and memory (Grandjean et al. 1997). Table 6.12 thus lists a

LOAEL for this study as less than 10 ppm maternal hair.

With the exception of the estimates for threshold for Minimata disease, the NOAEL, BMD and

threshold estimates (nine altogether) fall in a narrow range of 5-26.7 ppm maternal hair mercury. This

is a large degree of overlap with the 11ppm BMD and provides a great deal of support for this estimate

based on the Iraqi data. 

Chronic rodent (e.g. Bornhausen et al. 1980) and nonhuman primate studies (e.g. Burbacher et al.

1984; Gunderson et al. 1986; Rice et al. 1989a,b) provide data to support estimated NOAELs and 

LOAELs for developmental end points. The endpoints measured in these animal studies are relevant to

the types of toxicity which have been reported in children and they have been induced by dosing

protocols that are relevant to human exposures. Experiments in nonhuman primates have identified

adverse effects of methylmercury exposure in these areas: sensory (visual, somatosensory, auditory),

cognitive (learning under concurrent schedules, recognition of faces), social play, and

schedule-controlled operant behavior. The sensory, cognitive, and motor deficits appear reliable over a

consistent range of doses in nonhuman primates exposed to methylmercury during development. 
6-39

Table 6-13 lists NOAELs and LOAELs from animal studies. As the RfD for methylmercury is

based on effects in children who were exposed in utero, it is particularly useful to consider

developmental effects in animals exposed to methylmercury. NOAELs and LOAELs from selected

developmental studies of methylmercury in animals can be found in Table 6-14; a more complete

compilation is in section 3.3.3.8.
6-40

Table 6-13

Estimates of NOAELs and LOAELs from Animal Studies

Species/

Strain/

No. per NOAEL,

Sex per Exposure LOAEL

Group Duration (mg/kg-day) Effects Reference

Rat/Wistar/ 0–12 or NOAEL =2; Hindlimb crossing after 0–12 days Inouye and

10 F 12–20 d, LOAEL=4 Murakami

1x/d 1975

Rat/Wistar/ up to 26 NOAEL = Ruffled fur, loss of balance, hindlimb Munro et al.

50 F, 50 M mo 0.05; crossing, paralysis after 6 mo (males 1980

LOAEL=0.25 more affected); demyelination of dorsal

nerve roots and damage in teased

sciatic nerves at 0.25

Swiss 28 wk LOAEL = 1.9 Ataxia; degenerative changes of MacDonald

origin Purkinje cells; granule cell loss in and

Mouse cerebellum Harbison

M 1977

Cat/Breed 11 mo LOAEL = Degeneration of cerebellum and Chang et al.

NS/15-16 0.015 cerebral cortex; necrosis of dorsal root 1974

both sexes ganglia of kittens fed mercurycontaminated tuna

Cat/Breed 2 yr NOAEL = Impaired hopping reaction; decreased Charbonneau

NS/8-10 7 d/wk 0.020; LOAEL pain sensitivity; degeneration of dorsal et al. 1976

NS = 0.046 root ganglia 

Monkey/ 36–132 d NOAEL = Atrophy of neurons in calcarine cortex; Sato and

Macaca 1 x/d 0.02; LOAEL focal degeneration in sural nerves Ikuta 1975

fasciculari =, 0.03

s/1-2 both

sexes

Monkey/ 90-270 d Tremor; visual impairment Evans et al.

Macaca 1 x/wk NOAEL= 0.4; 1977

artoides, LOAEL = 0.5

Macaca

nemestrina

/2 both

sexes
Table 6-13 (continued)

Estimates of NOAELs and LOAELs from Animal Studies 

Species/

Strain/

No. per NOAEL,

Sex per Exposure LOAEL

Group Duration (mg/kg-day) Effects Reference

6-41

Monkey/ 4 yr LOAEL= 0.05 Spatial visual impairment Rice and

Macaca 7 d/wk Gilbert 1982

fasciculari 1 x/d

s/5

exposed, 2

control

(sex NS)

Monkey/ 3 yr LOAEL = 0.04 Slight tremor; motor incoordination; Burbacher et

Macaca 1 x/d blindness ; time to onset was 177-395 d al. 1988

fasciculari

s/7-8 F

Monkey/ 6.5-7 yr LOAEL = 0.05 Six years after end of dosing (follow-up Rice 1989b;

Macaca 7 d/wk study to Rice and Gilbert 1982): Rice and

fasciculari 1 x/d decreased fine motor performance; Gilbert 1992

s/4 M, 1 F diminished touch and pinprick

exposed, 1 sensitivity; impaired high frequency

M, 2 F hearing (p<0.05)

controls
6-42

Table 6-14

 NOAELs and LOAELs for Developmental Toxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex per Exposure NOAEL, LOAEL

Group Duration (mg/kg-day) Effects Reference

Rat/Holtzman during LOAEL=2.5 Decreased visual evoked potential Zenick 1976

/5 F gestation, latencies for peaks in 30-day old

during pups exposed during gestation,

lactation, during lactation, or during postnatal

or days 21-30.

postnatal

days 21-

30

Rat/Charles 47 d NOAEL=0.7, Ultrastructural changes, dose- Fowler and

River CD/ prior to LOAEL=1.4 related decrease in biochemical Woods 1977

20 F and activity in mitochondria of fetal

during hepatocytes 

gestation

Rat/Long- Once LOAEL=4 Increased P1-N1 amplitudes and Dyer et al.

Evans/4 Gd 7 decreased P2 and N2 latencies of 1978

exposed, 6 cortically visual evoked potential

control

Rat/Wistar- 4 d LOAEL=0.04 Impaired ability to perform operant Musch et al.

Neuherberg/ Gd 6-9 conditioning procedures (number 1978

No. F. NS of responses on lever required in

specified period of time) (

Rat/Wistar/10 4 d NOAEL=0.004, Reduction in behavioral Bornhausen et

F 1 x/d LOAEL=0.008 performance in offspring of treated al. 1980

Gd 6-9 mice following operant

conditioning

Rat/Sprague- 4 d NOAEL=1.6, Delayed vaginal patency, delayed Vorhees 1985

Dawley/ 1 x/d LOAEL=4.8 surface righting, retarded

15-19 F Gd 6-9 swimming development, lower

activity, impaired complex water

maze performance. Increased

mortality of pups at 1-21 days of

age
Table 6-14 (continued)

NOAELs and LOAELs from Developmental Toxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex per Exposure NOAEL, LOAEL

Group Duration (mg/kg-day) Effects Reference

6-43

Rat/Wistar/ during NOAEL=0.2, Increase in response latency in Schreiner et al.

38 M, 38 F gestation LOAEL=0.6 male (p<0.05) and female pups 1986

and (p<0.01) and in passiveness

lactation (p<0.05) in visual discrimination

reversal task 

Rat/HAN- 13 days LOAEL=0.2 Delayed sexual maturity (vaginal Suter and

Wistar/10 F prior to opening and testes descent) Schön 1986

mating

until

postnatal day

21

Rat/Wistar/16 2 wk LOAEL=0.08 Impaired tactile-kinesthetic Elsner 1991

F prior to function (p0.05) 

mating

through

weaning

Mouse/SvSl/ Once LOAEL=0.16 Impaired swimming ability and Spyker et al.

No. F NS Gd 7 or 9 open-field behavior (p<0.05) in 30- 1972

(i.p.) day old pups. 

Monkey/ approx. LOAEL=0.04 Impaired visual recognition Gunderson et

Macaca 1-3 yr 1 memory (data pooled from both al. 1988

fascicularis/9 x/d prior groups of infants of exposed

F exposed, 8 to mating mothers) compared to unexposed

F control through controls; test performed at 50-60

gestation days of age.

Monkey/ approx. 4 LOAEL=0.04 Decrease in social play behavior Gunderson et

Macaca mo to 2 and concomitant increase in al. 1988

fascicularis/ yr 1 x/d nonsocial passive behavior

12 F exposed, prior to compared to unexposed controls;

13 F control mating tests performed at 2 weeks to

through 8 months of age.

gestation 
Table 6-14 (continued)

NOAELs and LOAELs from Developmental Toxicity of Methylmercury in Animals

Species/

Strain/

No. per Sex per Exposure NOAEL, LOAEL

Group Duration (mg/kg-day) Effects Reference

6-44

Monkey/ 4-4.5 yr LOAEL=0.01 Spatial visual impairment Rice and

Macaca 1 x/d in Gilbert 1990

fascicularis/5 utero and

mothers postnatal

ly

Monkey/ 6.5-7 yr LOAEL=0.05 Six years after end of dosing Rice 1989b;

Macaca 7 d/wk (follow-up study to Rice and Rice and

fascicularis/4 1 x/d Gilbert 1982); decreased fine motor Gilbert 1992

M, 1 F performance; diminished touch and

exposed, 1 M, pinprick sensitivity; impaired high

2 F controls frequency hearing (p<0.05).

Monkey/ unspecifi NOAEL=0.08 No effect on spatial memory of Gilbert et al.

Macaca ed period adult offspring of animals treated 1993

fascicularis/ prior to with methylmercury hydroxide

23 F mating (data pooled from 24 animals, all

through treated groups).

gestation

Monkey/Saim week 11 LOAEL=0.7 to Monkeys exposed in utero tested Newland et al.

iri sciureus/3 or 14.5 0.9 ppm (on learned lever pulling activity) 1994

F until methylmercury at ages 5-6 yr. decreased

parturitio in maternal sensitivity to degrees in

n blood reinforcement; change in

reinforcement degree resulted in

either no behavior change or slow

change by comparison to controls.

At least three long term studies of non-human primates exposed in utero or as infants have been

undertaken. Description of the study populations (adapted from Rice 1996) is given in Table 6-15.

Postnatal, pre-pubescent exposure of the primates approximates the exposures experienced by

human children in both the short term (e.g. Iraq) and long term, steady state (e.g. Seychelles) scenarios

used in risk assessment. According to Guidelines for Risk Assessment established by the U.S. EPA

(1991), exposures of organisms until the time of sexual maturation should be considered in the

assessment of developmental toxicants. 

Exposure of monkeys to methylmercury during development has produced effects on sensory

systems; visual, auditory and somatosensory changes have been observed. These changes were
6-45

apparently permanent. There were inter-individual variations in the degree and type of impairment. 

These observations have implications for human studies in that they emphasize the necessity of assessing

the function integrity of multiple sensory systems. 

Rice (1996) noted that delayed neurotoxic effects were observed in monkeys six years after

exposure was stopped. This was expressed as clumsiness and decreased ability to grasp a variety of

objects when the animals were 13 years old. The author pointed out the relevance of this finding to the

potential for neurodegenerative disease development in aging humans exposed to methylmercury earlier

in life. There is increasing evidence, from Minamata and other exposures, that degenerative processes

associated with aging are exacerbated by mercury exposure (Igata 1993; Schantz et al. 1996).

Monkeys exposed to methylmercury in utero showed variable effects on cognitive processes

(Rice 1996). These animals were delayed in development of object permanence (ability to conceptualize

the existence of a hidden object) and recognition memory (recognition of faces). These abilities,

however, did not appear to be impaired in adult animals. The methylmercury effect appears to have been

a developmental delay rather than a permanent impairment. The emotional and social sequelae of a

similar reversible developmental delay in humans is not known. Newland et al. (1994) demonstrated a

persistent learning impairment in a group of three squirrel monkeys (Saimiri sciureus) exposed to

methylmercury in utero. They concluded that the toxic effect led to insensitivity to changes in the

consequences of behaviors. 
6-46

Table 6-15 

Studies in Non-Human Primates

Investigators /Exposure schedule dose n Total blood mercury in ppm

μg/kg/d

Rice and co-workers 0 5 peak at 200 Steady state

birth- 7 yrs <0.1

days <0.1

50 5 1.2 0.7

in utero - 4yrs 0 5 mothers birth Steady state

<0.1 <0.1 <0.1

10 1 0.3 0.5 0.2

25 2 0.7 0.9 0.4

50 5 1.4 2.7 0.7

Gunderson, Burbacher et al. 0 12 mothers birth

in utero

50 12 1.1 1.6

Newland et al 0 3 mothers birth

in utero

3 0.7 - 0.9

The study of visual psychophysics provides information about function and dysfunction in the

visual system, which can be applied to studies in humans. For example data on contrast sensitivity

functions in non-human primates provide links between important features of visual function as

expressed in behavior and the neural mechanisms underlying vision. The learning impairments

observed in behavior under concurrent schedules not only raise concerns about cognitive effects of

methylmercury exposure but also point to behavioral mechanisms by which these effects occur. Results

suggesting methylmercury-related deficits in the visual recognition of faces are congruent with

well-established areas of neuroscience that show how higher-order functioning is accomplished in the

primate sensory (including visual) cortex. These results point to links between the integration of

complex visual information and higher order cognitive abilities. 

The sensory and motor deficits observed in animals exposed to methylmercury indicate that the

exposed individual is missing the full complement of important capabilities. There are implications for

the adversity and long-term impact of some of the subtle changes in cognitive ability noted in children

exposed to mercury in utero. This research is leading to the recognition that forms of learning and

reading dysfunction in people can be traced to subtle alterations in sensory systems; these findings raise

concerns about deficits in functional domains not traditionally linked directly to sensory function. The

motor deficits are consistent with neural systems that are affected by methylmercury and, therefore,

indicate non-trivial impairment of the individual. Whether the cognitive endpoints are traceable to this

sensory loss remains to be determined, but some, such as the learning deficits under concurrent schedules

or alterations in fixed-interval schedule performance may be independent of such loss.
6-47

The rodent studies support conclusions drawn from non-human primates, although the results are

not always as consistent. Rats have most often been used, and the endpoints identified to date have

usually been less specific than those examined in the primate literature. Deficits in schedule-controlled

operant behavior (Bornhausen et al. 1980; Schreiner et al. 1986) and subtle characteristics of motor

function (Elsner 1991) have, however, been reported. The rodent studies generally find effects at doses 

predictable from a consideration of the kinetics of methylmercury in these species and the sensitivity of

the procedures used. Data on cognitive function generally show only weak effects and at high dose. 

Most often observed have been deficiencies in motor function. Sensory system function has not been as

extensively tested, although there have been reports of effects for in utero exposure on visual evoked

potentials (Zenick 1976; Dyer et al. 1978). 

Rice (1996) used data from animal studies described above to derive RfDs for methylmercury.

She identified LOAELs (in the absence of a NOAEL) of 0.01 to 0.05 mg/kg/day from the studies of

effects in monkeys exposed in utero and/or post partum. Standard U.S.EPA methodology was followed

in the application of 10-fold uncertainty factors for the following: variability in human populations,

extrapolation from animal data and use of a LOAEL in the absence of a NOAEL. Dividing by the

uncertainty factor of 1000 results in a range of RfDs of 0.01 to 0.05 μg/kg/day. If one uses the rat data, a

NOAEL of 0.005 mg/kg/day is identified and two 10 fold uncertainty factors are most appropriate (for

human variability and extrapolation from animal data). The resulting RfD would be 0.05 μg/kg/day. 

RfDs based on use of sensitive, but relevant, endpoints measured in animals are 5- to 10-fold lower than

those calculated from data in humans. This leads to a conclusion that the RfD based on observation of

clinical and other effects in Iraqi children is not unduly conservative. 

Uncertainty in the dose conversion (maternal hair mercury to dietary intake)

It was assumed in the derivation of the RfD that there was no substantial difference in

pharmacokinetics of methylmercury as a consequence of the food medium in which it is presented to

humans. The recent SAB report (U.S.EPA 1997) makes the following statement. 

There is no compelling evidence to suggest that the toxicokinetics of methylmercury (MeHg)

ingested in grain (as in the Iraqi poisoning episode) is different from that resulting form ingestion of fish

(the typical exposure route of humans). The best evidence for this is a study in which cats were fed

contaminated fish, control fish, or MeHg in a non-fish diet (Charbonneau et al., 1974). No differences

were observed in degree of MeHg neurotoxicity, latency to toxicity (ataxia), tissue levels or distribution

of Hg.

Gearhart et al. (1995) applied a PBPK model to their benchmark dose calculated in ppm maternal

hair mercury from the Kjellstrom et al. (1986a, b, 1989) data. Details of the application of this model

were not provided in the publication. A dose-conversion factor of about 0.05 [ppm mercury in

hair/(μg/kg-day)] can be estimated from Figure 5c in Gearhart et al. (1995). Applying this factor to the

benchmarks of 17 to 50 ppm mercury in hair yields the reported intake range of 0.85 to 25 μg/kg-day. 

Although this dose-conversion factor is about half of that estimated by the U.S. EPA, the two estimates

are consistent when duration of exposure is considered. The Gearhart et al. (1995) pharmacokinetic

model predicts that equilibrium is not reached until about 400 days. The dose conversion of 0.05

corresponds to hair mercury levels at equilibrium and is the appropriate factor to apply to the New

Zealand hair concentrations, which arose from longer-term exposure. In contrast, the Iraqi exposure was

only for a few months (Marsh et al. 1978). A dose conversion of about 0.1, which is virtually the same

as that used by the U.S. EPA, can be estimated for a 3-month exposure from Figure 4 in Gearhart et al.

(1995) and from Sherlock et al. (1984) assuming a hair:blood mercury concentration ratio of 250:1.

In their 1994 Toxicological Profile, ATSDR used the analysis reported by Cox et al. (1989), (see

discussion below) of the Iraqi developmental data in the derivation of an intermediate MRL (minimal
d  CxbxV

A x f x bw

 44 g/L x 0.013 days1

x 5L

0.95 x 0.077 x 60 kg

6-48

risk level). Using delayed onset of walking as the critical effect, a LOAEL of 14 ppm mercury in hair

was determined. A dose conversion from ppm hair to daily intake to maintain blood mercury levels in

pregnant women was done in a very similar manner to that employed by U.S. EPA. Values for

parameters in the equation on page 6-18 were consistent between the two Agencies with one exception;

namely the use of a blood volume of 4.1L by ATSDR compared to 5L by U.S. EPA. The methylmercury

intake level calculated by ATSDR to maintain a hair level of 14 ppm is 1.2 g/kg-day compared to 1.1

g/kg-day to maintain a hair level of 11 ppm (used by U.S. EPA).

ATSDR (1997) has recently released for public comment an updated Toxicological Profile with

a revised Minimal Risk Level (MRL) for methylmercury. In their dose conversion, they used the

following values: hair:blood ratio = 250; body weight = 60 kg; blood volume = 4.2 L; elimination

constant = 0.014; absorbed dose found in the blood = 0.05; absorption factor = 95%.

The state of New Jersey currently uses an RfD of 0.7x10 mg/kg-day (described in Stern 1993) -4

compared to the U.S. EPA's RfD of 1x10 mg/kg-day. The critical effect chosen was developmental -4

endpoints in the Iraqi children exposed in utero including delayed onset of walking. A recent discussion

of this RfD was presented in the context of the external peer review of the Mercury Study Report to

Congress. Stern described the LOAEL as the mercury hair level equivalent to a mercury blood level of

44 g/L. To determine the intake level, the equation on page 6-18 was used but with different values for

two parameters; namely, b and f.

= 0.70 g/kg-day

Choice of the value of 0.077 for f, fraction of daily intake taken up by blood was based on a

paper by Smith et al. (1994) which was not published at the time of the RfD/RfC work group discussions

and was not considered by them in choosing the parameter values. Smith et al. (1994) (described briefly

in chapter 2 of this volume) presents a study of methylmercury excretion kinetics based on measurement

of i.v. administered methylmercury (1.7-7.4 μg) in blood, urine and feces of 7 male volunteers. The

authors claim that data from this study are superior to those from previous studies in accounting for the

portion of the labeled mercury which is metabolized to inorganic mercury. Based on the linear

extrapolation of the plot of blood concentration of methylmercury versus time, the authors calculated that

approximately 7.5% of the methylmercury remained in the blood following rapid equilibration among

tissue compartments. Based on fitting the experimental data to a five compartment pharmacokinetic

model, they calculate that 7.7% (geometric mean) of the methylmercury is found in the blood. It should

be noted that the values for this parameter among the seven subjects ranged from 6.5-9.5%.

Smith et al. (1994) taking conversion to inorganic mercury into account, reported an overall

estimate (geometric mean) of the half-life in blood (methylmercury-specific as per discussion in previous

paragraph) of 45 days (0.015 days ). The half-lives (elimination constants) for the 7 subjects ranged -1

from 35 days (0.020 days ) to 53 days (0.013 days ). Stern (1993) notes the half-life reported by Cox et -1 -1

al. (1989) was 48 days with a range of about 18-37. This corresponds to a value for b of 0.0144 day . -1

The mean value is not reported by Cox, but a Monte Carlo simulation of the data estimated a mean of

about 47 days. The most frequently reported value (mode), however, was 55 days corresponding to a
6-49

value for b of 0.013 day . Ultimately the value of b = 0.013 day was chosen by Stern as the most -1 -1

"typical" value.

Uncertainty factors

The current RfD was derived by application of an uncertainty factor of 10. This was intended to

cover three areas of uncertainty: lack of data from a two-generation reproductive assay; variability in the

human population, in particular the wide variation in biological half-life of methylmercury and the

variation that occurs in the hair to blood ratio for mercury; and lack of data on long term sequelae of

developmental effects. There was no factor applied for sensitive subpopulations as data were obtained

from exposure to a sensitive human group; namely, the developing fetus. 

The interpretation by some risk assessors is that the effects noted in the Iraqi population exposed

to contaminated grain are not being seen at similar doses of methylmercury delivered in utero via

contaminated seafood. One assessment by a scientist at FDA is that the U.S. EPA RfD of 1.0x10-4

mg/kg-day for methylmercury is somewhat conservative and is certainly protective; a suggestion was

made that the uncertainty factor could be decreased to 3, resulting in a RfD of 3.0x10 mg/kg-day. -4

Stern (1997) did an analysis of the degree of uncertainty associated with interindividual

variability applied to the dose conversion from maternal hair mercury to ingested daily dose of

methylmercury. His conclusion was that a calculated ingested dose intended to be inclusive of 95-99%

of women 18-40 years old would be 0.1 to 0.3 μg/kg/day (by contrast to EPA’s calculated 1 μg/kg/day). 

His recommended uncertainty factor of three (for lack of reproductive data and data on long-term

sequelae) would result in a RfD of 0.01 to 0.03 μg/kg/day. 

ATSDR has recently released for public comment an updated Toxicological Profile with a

revised Minimal Risk Level (MRL) for methylmercury. This report chooses as a NOAEL the median

maternal hair mercury reported by Davidson et al. (1995) for the 29 month old Seychellois children

tested with the BSID and Bayley Infant Behavior record. The Toxicological Profile characterizes the

reported decrease in the male children’s activity level as not adverse and chooses use of a midpoint of all

measured maternal hair levels rather than the highest measure or median of the top quartile. An

uncertainty factor was not used to account for human variability. 

The recent review of the Mercury Study Report to Congress by the SAB recommended that EPA

consider information suggesting that the uncertainty factor applied to the Iraqi data be increased. Their

arguments are as follows.

For example, the Faeroe Islands data (also animal data) indicates that the fetal exposure may be

greater than maternal exposure. In this study fetal cord blood mercury levels averaged 80.2 ppb while

maternal blood levels were only 38.1 ppb. Animal data supports that the fetus may act as a sink for

mercury. The report extensively reviews blood mercury kinetics but has little to say about fetal brain

mercury levels. Although the data are slight there are indications from recent monkey studies that the

brain mercury half-life is very long (Vahter et al. 1995). The RfD for MeHg is based on results from an

acute exposure study while most MeHg exposure is thought to be long term. This may be an additional

reason to increase the uncertainty factor. There are also indications of age related changes where MeHg

may accelerate neurodegeneration associated with aging from human data (Igata, 1993) and animal data

(Rice 1989a; 1989b)). In evaluating neurotoxic effects from low exposures such as with methylmercury,

it must be remembered that few individuals may actually demonstrate clinical signs of disease but many

individuals may suffer subtle changes which can produce total population effects.

Risks among subpopulations
6-50

The recent review by the SAB of the Mercury Study Report to Congress discusses at some length

the likelihood that there are differential responses among human subgroups. That discussion is excerpted

below. 

Effect modification occurs when, at equal doses of a toxicant, adverse outcomes are observed in

some members of a population but not others. The affected individuals may possess one or more

distinctive characteristics such as age (stage of development), gender, social class, or certain premorbid

health factors (e.g., diabetes, liver disease, pulmonary dysfunction) or genetic predisposing factors that

are not well represented in unaffected members of the population.

Most studies of effects of environmental agents on child development have treated potential

effect modifiers as covariates or confounders in multiple linear regression models without interaction

terms models, or as matching variables in comparing so-called exposed and unexposed groups. 

Interactions are infrequently explored or are dismissed.

The animal literature, however, presents numerous examples of neurotoxicity enhancement or

buffering as a result of species, strain, drug, and physical and social environmental interactions. An

example from the methylmercury literature is the phenomenon of male susceptibility. Studies have

shown that the relative risk of perinatal morbidity and mortality is higher in human males (Abramowizc

and Barnett 1970; Naeye et al, 1971), including the risk of poor reproductive outcomes and postnatal

development due to fetal exposure to industrial pollutants (Scragg et al. 1977; McKeown-Eyssen et al.

1983). Males also have a higher rate of cognitive developmental disability in the general population

(Gross and Wilson 1974) and display more profound intellectual deficits as a result of cortical lesions

(Bornstein and Matazarro 1984; Inglis and Lawson 1981). The McKeown-Eyssen et al. (1983) study

reported gender-related differences including dose-related deficits in sensorimotor behaviors assessed in

the BSID and increased prevalence of abnormal muscle tone and deep tendon. The latter effect was not

associated with methylmercury dose in females. In the Iraqi poisoning more severe neurological effects

were observed in male children (Marsh et al. 1987). Animal experiments have also observed sex

differences in neurodevelopmental vulnerability. For example, in Sager et al. (1984) a single low dose

of methylmercury administered to neonatal mice resulted in mitotic arrest in cells of the granule layer of

the cerebellum only in males.

There is evidence that lifestyle factors such as the quality of the home environment and nutrition

play a role in the expression of developmental neurotoxicity. The literature on connections between

social factors and methylmercury toxicity is sparse. Studies on lead, however, have found greater

neurocognitive deficits in exposed individuals from the lowest socio-economic groups (e.g., Bellinger et

al. 1989; Dietrich et al. 1987; Harvey et al. 1984; Lansdown et al. 1986).

The positive nutritional factors of a seafood diet may be a factor in the greater delay in the onset

of the Minamata as compared to the Iraqi outbreaks. Early results from the Faroe Islands studies have

shown a positive association between cord blood methylmercury concentrations and birth weight

(Grandjean et al. 1992; Grandjean et al. 1995). The authors attribute the finding to the benefits of n-3

polyunsaturated fatty acids in a high seafood diet and to the benefits of breast-feeding which, in itself,

can lead to higher methylmercury intake by the infant.

The protective effects of genetic and environmental factors may be expressed in the studies of

Seychellois children (Davidson et al. 1995). At 19 and 29, months scores on the Psychomotor

Development Index (PDI) were very negatively skewed. The mean PDI scores at 19 and 29 months were

1.7 and 1.3 standard deviations above the United States means of 100 +/- 16 points respectively. That is,

means for the PDI recorded for Seychelles children are above what would be classified as accelerated

performance on these scales. The developmental health of this population is also reflected in the

small number of subjects attaining abnormal scores on the DDST-R by comparison to samples in the

United States. Only 3 out of 737 individual examinations or 0.4% were rated as abnormal (i.e., below
6-51

the 10th percentile for U.S. norms). Accelerated motor development has been noted in previous studies

of African cultures and is also observed in African-American infants under two years of age.

The conclusion of the SAB, however, was that the data regarding effect modification in human

epidemiologic studies of mercury poisoning are currently too meager to base separate estimates of

human health risks or establish different RFD’s for various subpopulations.

Other areas of uncertainty

Birth date uncertainty would have an impact on exposure uncertainty if correspondence of

exposure and gestation was estimated (Marsh et al. 1978) from birth date to any great extent. That is,

exposure may have occurred to a lesser extent (or not at all) than assumed during the critical period of

gestation. The result would be a lower exposure associated with the observation, depending on the width

of the critical time window during gestation and on the importance of duration of exposure in the

elicitation of the particular effect. If the exposure occurred after the critical period, any observation of an

effect would be attributed to causes other than methylmercury and be included in the background.

Several scientists have suggested that a developmental toxicity RfD is needed for

methylmercury. This may not be necessary, however, if the critical effect is developmental toxicity and

the uncertainty factors used to estimate the lifetime RfD do not involve an adjustment for less than

lifetime exposure nor lack of complete data base.
RfC  LOAEL mg Hg/m3

UF

 0.009 mg/m3

30

 0.0003 mg/m3

6-52

6.3.1.2 Inhalation Reference Concentrations (RfCs)

Elemental mercury

The U.S. EPA has determined an RfC of 3x10 mg/m for elemental mercury (U.S. EPA 1994). -4 3

The inhalation RfC is based on neurologic toxicity observed in several human occupational studies. The

observed neurologic changes included hand tremor, increases in memory disturbances and slight

subjective and objective evidence of autonomic dysfunction. Fawer et al. (1983) measured intention

tremor (tremors that occur at the initiation of voluntary movements) in workers exposed to a TWA

concentration of 0.026 mg/m over an average of 15.3 years. It was noted, however, that the tremors 3

may have resulted from intermittent exposures to concentrations higher than the TWA.

Piikivi and colleagues conducted several studies in chloralkali workers on electroencephalogram

(EEG) abnormalities (Piikivi and Toulonen 1989); subjective measures of memory disturbance and sleep

disorders and objective disturbances in psychological performance (Piikivi and Hanninen 1989); and

subjective and objective symptoms of autonomic dysfunction such as induced pulse rate variations and

blood pressure responses (Piikivi 1989). U.S. EPA extrapolated an occupational exposure level

associated with these neurological changes of 0.0250.030 mg/m from blood levels, based on a 3

conversion factor calculated by Roels et al. (1987). The LOAEL (0.025 mg/m adjusted to 0.009 mg/m 3 3

for continuous exposure of the general population) was divided by an uncertainty factor of 30 (10 to

protect sensitive individuals and for use of a LOAEL, and 3 for the lack of reproductive studies in the

database) to yield the RfC of 3x10 mg/m . -4 3

The RfC was, thus, calculated in the following way.

This reference concentration was reviewed and verified by the RfD/RfC Work Group and was

verified on April 19, 1990. It was released under a special action by U.S. EPA (Jarabek 1992, personal

communication).

Confidence in the critical study, the data base, and, thus in the RfC were rated "medium" by the

Work Group. Factors which were positive for confidence in the critical study were the use of a sufficient

number of human subjects, inclusion of appropriate controls, sufficient exposure duration and that the

LOAEL can be corroborated in other studies. It was noted, however, that for all but one of the studies,

exposure had to be extrapolated from blood mercury levels. The lack of human or multispecies

reproductive or developmental studies precluded higher confidence in the data base.

Inorganic (mercuric) mercury

Developmental toxicity (skeletal abnormalities and retarded growth) in mice (Selypes et al.

1984) and autoimmune disease in Brown-Norway rats (Bernaudin et al. 1981) have also been observed

following inhalation exposures. Due to the limitations of these inhalation studies and the inadequacy of

the remaining toxicologic and pharmacokinetic data bases, the RfD/RfC Work Group determined the

derivation of an RfC is not possible. The posting of this determination on IRIS is proceeding concurrent

with the finalization of this Report to Congress.
6-53

Methylmercury

No estimate of risk from inhalation of methylmercury has been done by U.S. EPA.

6.3.1.3 Estimation of Risk from Dermal Exposure

The dermal contribution of the different mercury species to the total systemic exposure of each

of these mercury species may be important for a full characterization of risk to the potentially exposed

human populations. Many of the necessary data needed for conducting a dermal risk assessment are

currently lacking or not well enough understood to assess systemic exposure and risks from dermal

exposure to mercury species.

For any of these mercury species to be a dermal health hazard they must be absorbed across the

skin (epidermis and dermis) and be systemically distributed to the affected critical organ systems

(kidneys or CNS) via the circulatory system. The percutaneous absorption for each mercury species is

dependent on skin-specific factors (e.g., skin thickness, hydration, age, condition, circulation, and

temperature) and compound-specific factors (e.g., lipophilicity, polarity, chemical structure, volatility,

and solubility), which are involved in determining the rate and amount of absorption by the cutaneous

route. Currently there are few known or agreed upon percutaneous absorption rates available for any of

the mercury species of interest. Some data on percutaneous absorption (Kp = Permeability coefficient)

for the mercuric forms of mercury in aqueous media have been reported in U.S. EPA (1992).

The media (aqueous, vapors or soil) where the mercury species are found must also be

considered in dermal risk assessments. Each medium has its own set of factors that impact the specific

percutaneous absorption rates for each of the mercury species. For example, mercury compounds found

in aqueous media are dependent on factors such as solubility in water and increased hydration of the

skin. Mercury compounds associated with soil must be assessed for binding to the particular soil type of

concern, the adhesion of the soil of concern to skin, the desorption of the mercury compound from the

soil, and the absorption of the mercury compound across the skin and into the circulatory system. Other

aspects that must be considered with a dermal assessment are binding or sequestration of the mercury

species at the site of exposure or closely nearby, and the metabolism of the mercury species in the skin

that may result in oxidation/reduction of the mercury species to other valence states (thereby, potentially

resulting in different critical effects than from the originally absorbed compound).

At present, many of the necessary mercury species/media factors have not been fully ascertained

and as a result credible dermal risk assessments cannot be accomplished at this time. A more extensive

discussion of dermal exposure assessment for risk assessment can be found in Dermal Exposure

Assessment: Principles and Applications (EPA/600/8-91/011B, January 1992). 
6-54

6.3.2 Developmental Effects

6.3.2.1 Elemental Mercury

Elemental mercury was judged to have sufficient animal data for developmental toxicity. The

two studies which contribute most to the level of concern are Danielsson et al. (1993) and Fredriksson et

al. (1992). Both of these studies are limited as a basis for an RfD by the small numbers of animals DT

tested and by the very few dose groups. A further limitation is lack of data on gender differences. No

RfD for elemental mercury is available at this time. DT

6.3.2.2 Inorganic Mercury

There are no data from human studies which are suitable for derivation of an RfD . Inspection DT

of available animal studies indicates that there are five reports of developmental effects of inorganic

mercury given orally. In all of these, exposure was by gavage to pregnant animals, and effects were

monitored in progeny. Three papers were reported as abstracts giving few experimental details.

Rizzo and Furst (1972) treated Long Evans rats (5/group) with a single gavage dose of 2 mg

Hg/kg as mercuric oxide on either day 5, 12, or 19 of gestation. Animals were sacrificed on day 20 or 21

of gestation. No effects of treatment on gestation day 12 or 19 were noted. According to the authors

treatment on day 5 resulted in a higher percentage of growth retardation and inhibition of eye formation,

but no statistical analyses were done.

In Gale (1974), pregnant Golden hamsters were administered 0, 2.5, 5, 16, 22, 32, 47, or 63 mg

Hg/kg-day mercuric acetate via gavage on gestation day 8. When the pregnant animals were sacrificed

on day 12 or 14, there was a significant increase in the incidence of abnormal fetuses including small,

retarded, or edematous (combined), and/or malformed fetuses. The NOAEL for developmental toxicity

was 5 mg mercuric chloride/kg or 2 mg Hg/kg. Maternal toxicity included dose-related weight loss,

diarrhea, slight tremor, somnolence, tubular necrosis and hepatocellular vacuolization, but insufficient

data were provided to allow determination of a LOAEL or NOAEL for maternal toxicity.

The advantage of the Gale (1974) study as a basis for quantitation of potential risk is that several

doses of inorganic mercury were tested; the spacing of the doses was adequate for identification of both a

LOAEL and NOAEL. It is not recommended, however, that the NOAEL in Gale (1974) serve as the

basis for an RfD . There were relatively few animals tested (decreasing the overall sensitivity of the DT

assay) and not all endpoints were thoroughly evaluated. The test compound was administered on only

one day of gestation, and there is some question as to the suitability of the golden hamster for

developmental assays. The data base for developmental effects, while generally supportive of the

LOAEL is not adequate to determine if the measured endpoints were the most sensitive for

developmental effects of inorganic mercury.

6.3.2.3 Methylmercury

Weight of evidence for developmental toxicity indicates that a developmental toxicity RfD is

appropriate for methylmercury. A separate RfD may not be necessary as the critical effect for the DT

lifetime RfD is developmental toxicity. The current RfD (1x10 mg/kg-day) was based on -4

developmental endpoints in offspring of women exposed during pregnancy; it may be taken as protective

against developmental toxicity. For less than chronic exposures it should be noted that the RfD is not DT

intended as a lifetime exposure value.

6.3.3 Germ Cell Mutagenicity
6-55

Data do not support the generation of quantitative estimates for germ cell mutagenicity for any

form of mercury.

6.3.4 Carcinogenic Effects

6.3.4.1 Elemental Mercury

Elemental mercury is categorized as D, unable to classify as to human carcinogenicity. A

quantitative estimate for carcinogenic effect is, thus, inappropriate.

6.3.4.2 Inorganic Mercury

Quantification of the potential carcinogenic effects of mercuric chloride (classified as C, possible

human carcinogen) was not done. No increase in tumor incidence was observed in a carcinogenicity

study in which white Swiss mice were given 0.95 mg Hg/kg-day as mercuric chloride in drinking water

(Schroeder and Mitchener 1975). No statement regarding carcinogenicity was reported in a 2-year

feeding study in which rats were administered mercuric acetate in the diet at doses of 0, 0.02, 0.1, 0.4,

1.7 and 6.9 mg Hg/kg-day (Fitzhugh et al. 1950).

The incidence of squamous cell papillomas of the forestomach and thyroid follicular cell

carcinomas from NTP (1993) was evaluated. No slope factor was based on the forestomach tumors

because this type of tumor is probably the result of irritation of the forestomach, cell death and epithelial

proliferation. The carcinogenic mechanism may be specific to irritation at the high doses used in the

bioassay; use of these tumors as a basis for human health assessment of low doses of inorganic mercury

is inappropriate.

Regarding the thyroid carcinomas, a variety of drugs, chemicals, and physiological perturbations

result in the development of thyroid follicular tumors in rodents. For a number of chemicals, the

mechanism of tumor development appears to be a secondary effect of longstanding hypersecretion of

thyroid-stimulating hormone by the pituitary (Capen and Martin 1989; McClain 1989). In the absence of

such long-term stimulatory effects, induction of thyroid follicular cell cancer by such chemicals usually

does not occur (Hill 1989). Use of the incidence of thyroid tumors from NTP (1993) in low dose

extrapolation is, thus, questionable.

6.3.4.3 Methylmercury

Quantification of the potential carcinogenic effects of methylmercury (classified as C, possible

human carcinogen) was not done. No increased incidence of tumors was seen in rats exposed to doses of

up to 0.34 mg Hg/kg-day for 130 weeks (Mitsumori et al. 1983, 1984) or in cats exposed to a diet

containing up to 0.176 mg Hg/kg-day for 2 years (Charbonneau et al. 1976).

No slope factor was calculated for methylmercury based on the incidence of renal epithelial

tumors in male mice. The two studies by Mitsumori et al. (1981, 1990) were limited by high mortality in

the high-dose males, the only group to exhibit a statistically significant increase in tumor incidence. The

study by Hirano et al. (1986) was not limited by survival problems, but the tumors were observed in

conjunction with nephrotoxicity and appear to be a high-dose phenomenon that may not be linear at low

doses. The tumors appeared to originate from focal hyperplasia of the tubular epithelium induced as a

reparative change. The hyperplasia was not observed in tubular epithelium that was undergoing early

degenerative changes; thus, the tumors may not occur where degenerative changes do not occur. The

appropriateness of deriving a quantitative risk estimate using the assumption of linearity at low doses

based on data for which a threshold may exist is questionable.
6-56

6.4 Risk Assessments Done By Other Groups

Quantitative estimates of hazards of oral exposure to methylmercury exposure have been

considered by the Food and Drug Administration, Agency for Toxic Substances and Disease Registry

(ATSDR), the Department of Energy and several State agencies. Several inhalation workplace exposure

limits are available in the United States and other countries.

6.4.1 Food and Drug Administration

In 1969, in response to the poisonings in Minamata Bay and Niigata, Japan, the U.S. FDA

proposed an administrative guideline of 0.5 ppm for mercury in fish and shellfish moving in interstate

commerce. This limit was converted to an action level in 1974 (Federal Register 39, 42738, December 6,

1974) and increased to 1.0 ppm in 1979 (Federal Register 44; 3990, January 19, 1979) in recognition that

exposure to mercury was less than originally considered. In 1984, the 1.0 ppm action level was

converted from a mercury standard to one based on methylmercury (Federal Register 49, November 19,

1984).

The action level takes into consideration the tolerable daily intake (TDI) for methylmercury, as

well as information on seafood consumption and associated exposure to methylmercury. The TDI is the

amount of methylmercury that can be consumed daily over a long period of time with a reasonable

certainty of no harm. U.S. FDA (and WHO) established a TDI based on a weekly tolerance of 0.3 mg of

total mercury per person, of which no more than 0.2 mg should be present as methylmercury. These

amounts are equivalent to 5 and 3.3 g, respectively, per kilogram of body weight. Using the values of

methylmercury, this tolerable level would correspond to approximately 230 g/week for a 70 kg person

or 33 g/person/day. The TDI was calculated from data developed in part by Swedish studies of

Japanese individuals poisoned in the episode of Niigata which resulted from the consumption of

contaminated fish and shellfish and the consideration of other studies of fish-eating populations.

Based on observations from the poisoning event later in Iraq, U.S. FDA has acknowledged that

the fetus may be more sensitive than adults to the effects of mercury (Federal Register 44: 3990, January

19, 179; Cordle and Tollefson, 1984, U.S. FDA Consumer, September, 1994). In recognition of these

concerns, U.S. FDA has provided advice to pregnant women and women of child-bearing age to limit

their consumption of fish known to have high levels of mercury (U.S. FDA Consumer, 1994). U.S. FDA

believes, however, that given existing patterns of fish consumption, few women (less than 1%) eating

such high mercury fish will experience slight reductions in the margin of safety. However, due to the

uncertainties associated with the Iraqi study, U.S. FDA has chosen not to use the Iraqi study as a basis

for revising its action level. Instead, the U.S. FDA has chosen to wait for findings of prospective studies

of fish-eating populations in the Seychelles Islands and in the Faroes Islands.
6-57

6.4.2 ATSDR

ATSDR has established Minimal Risk Levels (MRLs) for elemental, inorganic and

methylmercury (ATSDR 1994). Recently a revised Toxicological Profile has been released for public

comment (ATSDR 1997).

An acute inhalation MRL of 0.00002 mg/m has been derived for elemental mercury vapor based 3

on neurodevelopmental changes in rats. Specifically, the effects were changes in locomotor activity at 4

months of age and an increased time to complete a radial arm maze at 6 months of age following

exposure to 0.05 mg Hg/m for 1 hour during post-partum days 1117 (Fredriksson et al. 1992). A 3

chronic inhalation MRL of 0.000014 mg/m was derived for elemental mercury vapor based on a 3

significant increased in the average velocity of naturally occurring tremors in occupational workers

(Fawer et al. 1983). The revised chronic MRL is calculated to be 0.0002 mg/m by application of an 3

uncertainty factor of 30 to a LOAEL of 0.026 mg/m for increased frequency of tremors in 3

occupationally exposed workers (Fawer et al. 1983).

Acute and intermediate oral MRLs were derived for inorganic mercury based on kidney effects

reported in the 1993 NTP study of mercuric chloride. The acute oral MRL was 0.007 mg Hg/kg-day

based on a 2-week study reporting a NOAEL of 0.93 mg Hg/kg-day for renal effects in rats (NTP 1993). 

At higher doses, an increased incidence of tubular necrosis was observed. The intermediate oral MRL of

0.002 mg Hg/kg-day was established, based on a 6-month study reporting a NOAEL of 0.23 mg Hg/kgday for renal effects (increased absolute and relative kidney weights) (NTP 1993). There is no indication

that these values have been revised in the 1997 document.

An acute-intermediate oral MRL of 0.00012 mg Hg/kg-day was established in 1994 for

methylmercury. ATSDR derived their assessment from the Marsh et al. (1981) and Cox et al. (1989)

data; the MRL is based on the lowest observed peak of total mercury concentration in maternal hair

(0.0012 mg/kg-day equivalent to a LOAEL of 14 ppm mercury in maternal hair) during pregnancy

associated with a delayed onset of walking in offspring in Iraqi children. This assessment is discussed in

section 6.3.1.1 of this volume.

The 1997 Toxicological Profile calculates a chronic MRL for methylmercury of 0.5 μg/kg/day. 

This report chooses as a NOAEL the median maternal hair mercury of 5.9 ppm reported by Davidson et

al. (1995) for the 29 month old Seychellois children tested with the BSID and Bayley Infant Behavior

record. The Toxicological Profile characterizes the reported decrease in the male children’s activity level

as not adverse and chooses use of a midpoint of all measured maternal hair levels rather than the highest

measure or median of the top quartile. Dose conversion was done as in the 1994 document to give an

estimated ingested dose of 0.5 μg/kg/day. An uncertainty factor was not used to account for human

variability. 

6.4.3 Department of Energy

Brookhaven Laboratories has prepared a report for Office of Clean Coal Technology, DOE. This

report describes a probabilistic-based assessment which considered the potential increased health risk for

paresthesia in adults. Their estimate is based upon a yearly emission rate of 180 kg/year from all fossil

fuel power plants in the United States. This estimate represents less than 1% of the existing global pool

of mercury that is introduced into the environment. Based upon the most sensitive adult sign of

paresthesia, the mercury emissions from power plants would result in an increased risk for paresthesia of

0.0040.007% with an upper 95th percentile risk of 0.0130.017% (Lipfert et al. 1994).
6-58

6.4.4 National Institute of Environmental Health Sciences (NIEHS)

NIEHS, part of the National Institutes of Health, was required under section 301 of the CAA "to

conduct, and transmit to the Congress by November 15, 1993, a study to determine the threshold level of

mercury exposure below which adverse human health effects are not expected to occur." In section 112

(n)(1)(C), NIEHS was encouraged to evaluate the health effects threshold for mercury in the absence of

specifics as to species of mercury but to consider mercury in fish. As mercury in fish is primarily in the

form of methylmercury, the NIEHS limited their consideration to this species.

The report was completed in 1993 and delivered to Office of Management and Budget for

clearance. It describes dose- response assessments for methylmercury done by WHO, FDA and U.S.

EPA and presents all three estimates as recommended for tolerable mercury concentrations. The NIEHS

report also describes estimates of fish consumption by the U.S. population.

6.4.5 Department of Labor

OSHA established a Permissible Exposure Limit (PEL), time-weighted average of 0.05 mg

Hg/m for mercury vapor, with a notation for skin exposure (U.S. Department of Labor 1989). A PEL as 3

a ceiling value of 0.1 mg Hg/m , also with a notation for dermal exposure was set for aryl mercury and 3

inorganic mercury compounds.

NIOSH determined a Recommended Exposure Limit (REL), time-weighted average, of 0.05 mg

Hg/m for mercury and 0.1 mg Hg/m for aryl and inorganic mercury compounds (NIOSH 1973, 1988). 3 3

6.4.6 Various States

A number of states have released fish consumption advisories based upon their independent

analysis of the available scientific literature for methylmercury. Most active among these states are

Michigan, New Jersey, Maine, Idaho, and Oregon. Generally, there is a trend to move to more

conservative values based upon developmental neurotoxicity defined in the Marsh et al. (1981) and Cox

et al. (1989) papers. The methylmercury RfD of 0.7x10 mg/kg-day used by the state of New Jersey is -4

discussed in section 6.3.1.1. Some states are waiting for more specific guidance from U.S. EPA.

6.4.7 World Health Organization

The International Programme on Chemical Safety (IPCS) of the World Health Organization

published a criteria document on mercury (WHO 1990). In that document, it was stated that " a daily

intake of 3 to 7 μg Hg/kg body weight would cause adverse effects of the nervous system, manifested as

an approximately 5% increase in the incidence of paraesthesias". The IPCS expert group also concluded

that developmental effects in offspring (motor retardation or signs of CNS toxicity) could be detected as

increases over background incidence at maternal hair levels of 1020 ppm mercury. These levels of

concern were based on evaluation of data including the human poisoning incident in Iraq described in

Chapter 3.
6-59

6.4.8 ACGIH

The ACGIH has established Threshold Limit values (TLV) as eight-hour time-weighted

averages. They include the following:

Aryl mercury compounds 0.1 mg Hg/m3

Mercury vapor 0.05 mg Hg/m3

Inorganic mercury 0.1 mg Hg/m3

No STEL is recommended at this time. The Biological Exposure Indices Committee has

recommended values for inorganic mercury in urine and blood of 35 μg/g creatinine and 15 μg/L

respectively.

The ACGIH classified inorganic mercury including elemental mercury as follows: A4- Not

classifiable as a Human Carcinogen: There are inadequate data on which to classify the agent in terms of

its carcinogenicity in humans and/or animals.
7-1

7. ONGOING RESEARCH AND RESEARCH NEEDS

7.1 Ongoing Research

Table 7-1 lists ongoing research projects abstracted from the Federal Research in Progress Data

Base (FEDRIP, 1994).

Table 7-1

Ongoing Research

Investigator Affiliation Research Description Sponsor

Human

T. Clarkson University of Rochester, Rochester, Dose-response relationships in humans National Institute of

NY exposed to methylmercury and prenatal Environmental Health

and early postnatal body burdens of Sciences (NIEHS)

methylmercury.

P. Grandjean Odense University, Odense, Neurotoxicity risk from exposure to NIEHS

Denmark methylmercury from seafood

W. Markesbery University of Kentucky, Lexington, Role of mercury and dental amalgams in National Institute on

KY Alzheimer's disease Aging

M. Martin University of Washington, Seattle, Epidemiology of mercury in dentists National Institute of

WA Dental Research

R. Mitchell University of Kentucky, Lexington, Amalgam restorations and the relative National Institute of

KY risk of adverse pregnancy outcome Dental Research

G. Myers University of Rochester, Rochester, Child development following prenatal NIEHS

NY methylmercury exposure via fish

T. Okabe Baylor College of Dentistry, Dallas, Establish maximum levels of exposure National Institute of

TX from amalgams for dental patients and Dental Research

personnel

M. Owens Science Applications International Potential and adverse effects associated National Institute of

Corp, with dental amalgam Dental Research

Falls Church, VA

M. Rosenman Morehouse College, Atlanta, GA Effect of mercury in amalgam and urine National Institute of

to cognitive functioning in children General Medical

Sciences

D. Savitz University of North Carolina Mercury and reproductive health in National Institute of

Chapel Hill, Chapel Hill, NC women dentists Dental Research

Animal

P. Bigazzi University of Connecticut, Mercury induced auto-immune disease in NIEHS

Farmington, CT rats

T. Burbacher University of Washington, Seattle, Developmental effects of methylmercury NIEHS

WA in monkeys and rats

K. Mottet University of Washington, Seattle, Long-term toxicity associated with NIEHS

WA inorganic mercury and methylmercury 

K. Pollard University of California, San Diego, Animal model of systemic autoimmunity National Institute of

CA induced by mercury Arthritis and

Musculoskeletal and

Skin Diseases
Table 7-1

Ongoing Research (continued)

Investigator Affiliation Research Description Sponsor

7-2

B. Weiss University of Rochester, Rochester, Neurotoxicity throughout the lifespan of NIEHS

NY mice exposed prenatally to

methylmercury

Mechanistic

W. Atchison Michigan State University, East Neurotoxic mechanism of chronic NIEHS

Lansing, MI methylmercury poisoning

D. Barfuss Georgia State University, Atlanta, Transport and toxicity of inorganic NIEHS

GA mercury in the nephron

T. Jensen Herbert H. Lehman College, New Effect on membrane structure and National Institute of

York, NY organelle distribution General Medical

Sciences

D. Lawrence Albany Medical College, Albany, Effects of metals on the structure and NIEHS

NY function of murine and human

lymphocytes

R. Noelle Dartmouth Medical School, Effect of mercury on -lymphocyte NIEHS

Hanover, NH function

K. Pollard Scripps Research Institute, San Mechanisms of autoantibody response National Institute of

Diego, CA induced by mercury which target the Allergy and Infectious

nucleolus Diseases

B. Rajanna Selma University, Biomechanisms of heavy metal toxicity in National Institute of

Selma, AL rats General Medical

Sciences

K. Ruehl Rutgers University, Mechanism of methylmercury NIEHS

New Brunswick, NJ neurotoxicity during development in mice

T. Sarafian University of California, Los Effect of methylmercury on protein NIEHS

Angeles, CA phosphorylation in cerebellar granule cells

in brain

J. Stokes Mount Desert Island Biological Effects of mercurials on transport NIEHS

Lab, properties of the bladder

Salsbury Cove, ME

R. Zalups Mercer University School of Cytotoxicity of mercuric chloride to NIEHS

Medicine isolated rat proximal tubular cells

Two of these ongoing studies deserve further discussion because they may fill critical data needs

for the development of a reference dose for methylmercury. The first is the Seychelles Islands Study led

by Dr. T.W. Clarkson from the University of Rochester. The objective of this study is to define the

extent of human health risks from prenatal exposure to methylmercury. Dose-response relationships in a

human population with dietary exposure to methylmercury at levels believed to be in the range of the

threshold for developmental toxicity are being studied. Both prenatal and early postnatal body burdens

of methylmercury will be examined as well as transport to the brain.

This study is testing the hypothesis, developed in previous studies of prenatal exposure in the

Iraq population, that subtle psychological and behavioral changes in prenatally exposed children can be

quantitatively related using dose-response models to the mother's methylmercury exposure during

pregnancy. In the Seychelles, a group of islands off the coast of Africa near Madagascar, a group of 779

infants who were prenatally exposed to methylmercury through maternal fish consumption is being

studied with annual administration of neurodevelopmental, psychological and educational testing of the
7-3

children through 5.5 years of age. This population consumes a relatively large amount of marine fish and

marine mammals, both of which are likely to contain methylmercury. The study is testing the hypothesis

that methylmercury concentration in hair correlates with methylmercury in the brain by using human

autopsy data. Mechanisms of transport of methylmercury across the blood brain barrier also are being

studied to understand better the factors that limit the accuracy of hair mercury as a biological marker for

target tissue levels. Findings reported in recent publications are summarized in section 3.3.1.1.

The second study is the Faroe Islands Study led by Dr. P.A. Grandjean from Odense University

in Denmark. The purpose of this study is to determine whether a neurotoxic risk is present from

methylmercury exposure from seafood and, if so, the threshold for such effects. This study is examining

a cohort of 1,000 children in the Faroe Islands, located in the North Atlantic between Scotland and

Iceland. As is the case in the Seychelles, this population consumes a relatively large amount of seafood;

consumption includes marine fish and marine mammals. Intrauterine exposures were determined by

mercury analysis of umbilical cord blood and maternal hair collected at consecutive births during 21

months in 1986 and 1987. In 13 percent of the births, mercury levels were greater than 10 ppm in

maternal hair, and 25 percent of the cord blood samples had a mercury concentration above the

corresponding level of 40 μg/L. No cases of gross methylmercury poisoning have been observed. The

persistence of mercury in the body is being assessed from mercury hair concentrations in the children at

one and six years of age, and dietary information is being collected. A detailed pediatric examination

and a test battery to identify possible subtle signs of neurobehavioral dysfunction are being conducted. 

The test battery includes psychological tests and neurophysiological measurement of evoked potentials;

these methods are known from previous research to be particularly sensitive to the types of neurotoxicity

expected.

The Faroese population was chosen for this study because of the homogeneity and stability of the

population and the efficient coverage of the Danish health care system. The cohort includes 75% of all

births occurring during the sampling period. A high participation rate (about 80%) is expected at the 6-

year examination period. Alcohol use is minimal in Faroese women (75% were abstainers during

pregnancy), and 60% are nonsmokers. The lead exposure is low (median lead concentration in cord

blood was 1.7 μg/100 mL). Exposure to polychlorinated biphenyls (PCB), however, may be a

confounder, and alcohol intake of the fathers may have been high. Due to the high seafood intake,

selenium exposure is increased, and its possible protective action against mercury toxicity is being

examined. Findings reported at recent scientific meetings are summarized in section 3.3.1.1.

7.2 Research Needs

In addition to the ongoing studies described above, further research is necessary for refinement

of the U.S. EPA's risk assessments for mercury and mercury compounds. In order to reduce uncertainties

in the current estimates of the oral reference doses (RfDs) and inhalation reference concentrations

(RfCs), longer-term studies with low-dose exposures are necessary. In particular, epidemiological

studies should emphasize comprehensive exposure data with respect to both dose and duration of

exposure. The current RfD and RfC values have been determined for the most sensitive toxicity endpoint

for each compound; that is, the neurological effects observed following exposure to elemental or

methylmercury, and the renal autoimmune glomerulonephritis following exposure to inorganic mercury. 

For each of these compounds, experiments conducted at increasingly lower doses with more sensitive

measures of effect will improve understanding of the respective dose-response relationships at lower

exposure levels and the anticipated thresholds for the respective effects in humans. Similar information

from developmental toxicity studies would allow determination of RfDs for developmental toxicity

(RfD ) for elemental and inor dt ganic mercury. For inorganic mercury, furthermore, the many ongoing

studies in which mechanisms of action are being investigated will greatly assist in quantifying the risks

posed by these compounds.
7-4

Well-conducted studies are also needed to clarify exposure levels at which toxic effects other

than those defined as “critical” could occur in humans. For all three forms of mercury, data are

inadequate, conflicting, or absent for the following: adverse reproductive effects (effects on function or

outcome, including multigeneration exposure); impairment of immune function; and genotoxic effects on

human somatic or germinal cells (elemental and inorganic mercury). Investigations that relate the toxic

effects to biomonitoring data will be invaluable in quantifying the risks posed by these mercury

compounds. In addition, work should focus on subpopulations that have elevated risk because they are

exposed to higher levels of mercury at home or in the workplace, because they are also simultaneously

exposed to other hazardous chemicals, or because they have an increased sensitivity to mercury toxicity. 

Information on postnatal exposure without prenatal exposure is limited; therefore, analyzing the potential

risks associated with mercury exposure of young children is difficult.

There are data gaps in the carcinogenicity assessments for each of the mercury compounds. The

U.S. EPA's weight-of-evidence classification of elemental mercury (Group D) is based on studies in

workers who were also potentially exposed to other hazardous compounds including radioactive isotopes,

asbestos, or arsenic. There were no appropriate animal studies available for this compound.

Studies providing information on the mode of action of inorganic mercury and methylmercury in

producing tumors will be of particular use in defining the nature of the dose response relationship.

The assessment of both noncarcinogenic effects and carcinogenic effects will be improved by an

increased understanding of the toxicokinetics of these mercury compounds. In particular, quantitative

studies that compare the three forms of mercury across species and/or across routes of exposure are vital

for the extrapolation of animal data when assessing human risk. For elemental mercury there is a need

for quantitative assessment of the relationship between inhaled concentration and delivery to the brain or

fetus; in particular the rate of elemental to mercuric conversion mediated by catalase and the effect of

blood flow. Such assessment is needed for evaluation of the impact of mercury exposure from dental

amalgam.

Work has been done on development of physiologically-based pharmacokinetic models. While

one of these has developed a fetal submodel, data on fetal pharmacokinetics are generally lacking. The

toxicokinetics of mercury as a function of various developmental stages should be explored. Elemental

mercury and methylmercury appear to have the same site of action in adults; research is, therefore,

needed on the potential for neurotoxicity in newborns when the mother is exposed. This work should be

accompanied by pharmacokinetic studies and model development.
8-1

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APPENDIX A

DOSE CONVERSIONS
A-1

APPENDIX A

DOSE CONVERSIONS

All doses in the tables in Section 4 were adjusted for the amount of mercury in the compound.

For example, for animals administered 1 mg/kg/day mercuric chloride:

Molecular weight of mercuric chloride = 271.5

Molecular weight of mercury = 200.6

Dose of mercury = 1 mg Hg/kg/day x 200.6/271.5 = 0.74 mg Hg/kg/day

(1) To convert from ppm in feed to mg/kg body weight/day, the following equation was used:

mg toxicant (T)/kg body weight/day = mg T/kg food x food factor

where food factor = kg food per day intake/kg body weight

(2) To convert from ppm in water to mg/kg body weight/day, the following equation was used:

mg T/kg body weight/day = mg T/L water x L water per day intake/kg body weight

where L is liters of water intake per day

Species

Sample Values Used

Water Food factor

intake/day Body weight (kg food/kg body

(Liter/day) (kg) weight)

Mouse 0.0057 0.03 0.13

Rat 0.049 0.35 0.05

Rabbit 0.41 3.8 0.049

(3) To convert from ppm in air to mg/m for a vapor, the following equation was used: 3

1 mg/m = 1 ppm x molecular weight/24.45 3
APPENDIX B

SUMMARIES FOR THE

INTEGRATED RISK INFORMATION SYSTEM (IRIS)
B-1

 I.B. REFERENCE CONCENTRATION FOR CHRONIC INHALATION EXPOSURE (RfC)

Substance Name -- Elemental mercury (Hg)

CASRN -- 7439-97-6

Preparation date -- 3/12/90

 I.B.1. INHALATION RfC SUMMARY

Critical Effect Exposures* UF MF RfC

Hand tremor; increases NOAEL: None 30 1 3E-4

in memory disturbances; mg/cu.m

slight subjective and LOAEL: 0.025 mg/cu.m

objective evidence of (converted to LOAEL [ADJ]

autonomic dysfunction of 0.009 mg/cu.m

Human occupational

inhalation studies

Fawer et al., 1983;

Piikivi and Tolonen, 1989;

Piikivi and Hanninen, 1989;

Piikivi, 1989;

Ngim et al., 1992;

Liang et al., 1993

* Conversion Factors and Assumptions: This is an extrarespiratory effect of a vapor (gas). The LOAEL

is based on an 8-hour TWA occupational exposure. MVho = 10 cu.m/day, MVh = 20 cu.m/day. 

LOAEL(HEC) = LOAEL(ADJ) = 0.025 mg/cu.m x MVho/MVh x 5 days/7 days = 0.009 mg/cu.m. Air

concentrations (TWA) were measured in the Fawer et al. (1983), Ngim et al. (1992), and Liang et al.

(1993) studies. Air concentrations were extrapolated from blood levels based on the conversion factor of

Roels et al. (1987) as described in the Additional Comments section for the studies of Piikivi and

Tolonen (1989), Piikivi and Hanninen (1989), and Piikivi (1989).

 I.B.2. PRINCIPAL AND SUPPORTING STUDIES (INHALATION RfC)

Fawer, R.F., U. DeRibaupierre, M.P. Guillemin, M. Berode and M. Lobe. 1983. Measurement of hand

tremor induced by industrial exposure to metallic mercury. J. Ind. Med. 40: 204-208. 

Piikivi, L. and U. Tolonen. 1989. EEG findings in chlor-alkali workers subjected to low long term

exposure to mercury vapor. Br. J. Ind. Med. 46: 370-375. 

Piikivi, L. and H. Hanninen. 1989. Subjective symptoms and psychological performance of

chlorine-alkali workers. Scand. J. Work Environ. Health. 15: 69-74. 
B-2

Piikivi, L. 1989. Cardiovascular reflexes and low long-term exposure to mercury vapor. Int. Arch.

Occup. Environ. Health. 61: 391-395. 

Ngim, C.H., S.C. Foo, K.W. Boey and J. Jeyaratnam. 1992. Chronic neurobehavioral effects of

elemental mercury in dentists. Br. J. Ind. Med. 49: 782-790. 

Liang, Y-X., R-K. Sun, Y. Sun, Z-Q. Chen and L-H. Li. 1993. Psychological effects of low exposure to

mercury vapor: Application of a computer-administered neurobehavioral evaluation system. Environ.

Res. 60: 320-327. 

Fawer et al. (1983) used a sensitive objective electronic measure of intention tremor (tremors that

occur at the initiation of voluntary movements) in 26 male workers (mean age of 44 years) exposed to

low levels of mercury vapor in various occupations: fluorescent tube manufacture (n=7), chloralkali

plants (n=12), and acetaldehyde production (n=7). Controls (n=25; mean age of 44.6 years) came from

the same factories but were not exposed occupationally. Personal air samples (two per subject) were

used to characterize an average exposure concentration of 0.026 mg/cu.m. It should be noted that it is

likely that the levels of mercury in the air varied during the period of exposure and historical data

indicate that previous exposures may have been higher. Exposure measurements for the control cohort

were not performed. The average duration of exposure was 15.3 years. The measures of tremor were

significantly increased in the exposed compared to control cohorts, and were shown to correspond to

exposure and not to chronologic age. These findings are consistent with neurophysiological impairments

that might result from accumulation of mercury in the cerebellum and basal ganglia. Thus, the TWA of

0.026 mg/cu.m was designated a LOAEL. Using the TWA and adjusting for occupational ventilation

rates and workweek, the resultant LOAEL(HEC) is 0.009 mg/cu.m. 

Piikivi and Tolonen (1989) used EEGs to study the effects of long-term exposure to mercury

vapor in 41 chloralkali workers exposed for a mean of 15.6 +/- 8.9 years as compared with matched

referent controls. They found that the exposed workers, who had mean blood Hg levels of 12 ug/L and

mean urine Hg levels of 20 ug/L, tended to have an increased number of EEG abnormalities when

analyzed by visual inspection only. When the EEGs were analyzed by computer, however, the exposed

workers were found to have significantly slower and attenuated brain activity as compared with the

referents. These changes were observed in 15% of the exposed workers. The frequency of these changes

correlated with cortical Hg content (measured in other studies); the changes were most prominent in the

occipital cortex less prominent in the parietal cortex, and almost absent in the frontal cortex. The authors

extrapolated an exposure level associated with these EEG changes of 0.025 mg/cu.m from blood levels

based on the conversion factor calculated by Roels et al. (1987). 

Piikivi and Hanninen (1989) studied the subjective symptoms and psychological performances

on a computer-administered test battery in 60 chloralkali workers exposed to mercury vapor for a mean

of 13.7 +/- 5.5 years as compared with matched referent controls. The exposed workers had mean blood

Hg levels of 10 ug/L and mean urine Hg levels of 17 ug/L. A 

statistically significant increase in subjective measures of memory disturbance and sleep disorders was

found in the exposed workers. The exposed workers also reported more anger, fatigue and confusion. 

No objective disturbances in perceptual motor, memory or learning abilities were found in the exposed

workers. The authors extrapolated an exposure level associated 

with these subjective measures of memory disturbance of 0.025 mg/cu.m from blood levels based on the

conversion factor calculated by Roels et al. (1987).
B-3

Both subjective and objective symptoms of autonomic dysfunction were investigated in 41

chloralkali workers exposed to mercury vapor for a mean of 15.6 +/- 8.9 years as compared with matched

referent controls (Piikivi, 1989). The quantitative non-invasive test battery consisted of measurements of

pulse rate variation in normal and deep breathing, in the Valsalva maneuver and in vertical tilt, as well as

blood pressure responses during standing and isometric work. The exposed workers had mean blood

levels of 11.6 ug/L and mean urine levels of 19.3 ug/L. The exposed workers complained of more

subjective symptoms of autonomic dysfunction than the controls, but the only statistically significant

difference was an increased reporting of palpitations in the exposed workers. The quantitative tests

revealed a slight decrease in pulse rate variations, indicative of autonomic reflex dysfunction, in the

exposed workers. The authors extrapolated an exposure level associated with these subjective and

objective measures of autonomic dysfunction of 0.030 mg/cu.m from blood levels based on the

conversion factor calculated by Roels et al. (1987).

Two more recent studies in other working populations corroborate the neurobehavioral toxicity

of low-level mercury exposures observed in the Fawer et al. (1983), Piikivi and Tolonen (1989), Piikivi

and Hanninen (1989), and Piikivi (1989) studies.

Ngim et al. (1992) assessed neurobehavioral performance in a cross-sectional study of 98 dentists

(38 female, 60 male; mean age 32, range 24-49 years) exposed to TWA concentrations of 0.014 mg/cu.m

(range 0.0007 to 0.042 mg/cu.m) versus 54 controls (27 female, 27 male; mean age 34, range 23-50

years) with no history of occupational exposure to mercury. Air concentrations were measured with

personal sampling badges over typical working hours (8-10 hours) and converted to an 8-hour TWA. No

details on the number of exposure samples or exposure histories were provided. Blood samples from the

exposed cohort were also taken and the data supported the correspondence calculated by Roels et al.

(1987). Based on extrapolation of the average blood mercury concentration (9.8 ug/L), the average

exposure concentration would be estimated at 0.023 mg/cu.m. The average duration of practice of the

exposed dentists was 5.5 years. Exposure measurements of the control cohort were not performed. The

exposed and control groups were adequately matched for age, amount of fish consumption, and number

of amalgam dental fillings. The performance of the dentists was significantly worse than controls on a

number of neurobehavioural tests measuring motor speed (finger tapping), visual scaning, visumotor

coordination and concentration, visual memmory, and visuomotor coordination speed. These

neurobehavioral effects are consistent with central and peripheral neurotoxicity and the TWA is

considered a LOAEL. Using the TWA and adjusting for occupational ventilation rates and the reported

6-day workweek, the resultant LOAEL(HEC) is 0.006 mg/cu.m. 

Liang et al. (1993) investigated workers in a fluorescent lamp factory with a

computer-adminstered neurobehavioral evaluation system and a mood inventory profile. The exposed

cohort (mean age 34.2 years) consisted of 19 females and 69 males exposed to ninterruptedly for at least

2 years prior to the study. Exposure was monitored with area samplers and ranged from 0.008 to 0.085

mg/cu.m across worksites. No details on how the exposure profiles to account for time spent in different

worksites were constructed. The average exposure was estimated at 0.033 mg/cu.m. (range 0.005 to 0.19

mg/cu.m). The average duration was of working was 15.8 years for the exposed cohort. Urinary

excretion was also monitored and reported to average 0.025 mg/L. The control cohort (mean age 35.1

years) consisted of 24 females and 46 males recruited from an embroidery factory. The controls were

matched for age, education, smoking and drinking habits. Exposure measurements for the control cohort

were not performed. The exposed cohort performed significantly worse than the control on tests of

finger tapping, mental arithmetic, two-digit searches, switiching attention, and visual reaction time. The

effect on performance persisted after the confounding factor of chronological age was controlled. Based
B-4

on these neurobehavioral effects, the TWA of 0.033 mg/cu.m is designated as LOAEL. Using the TWA

and adjusting for occupational ventilation rates and workweek, the resultant LOAEL(HEC) is 0.012

mg/cu.m. 

The above studies were taken together as evidence for a LOAEL based on neurobehavioral

effects of low-level mercury exposures. The LOAEL(HEC) levels calculated on measured air

concentration levels of the Ngim et al. (1992) and the Liang et al. (1993) studies bracket that calculated

based on the air concentrations measured by Fawer et al. (1983) as a median HEC level. Extrapolations

of blood levels, used as biological monitoring that accounts for variability in exposure levels, also

converge at 0.025 mg/cu.m as a TWA which results in the same HEC level. Thus, the TWA level of

0.025 mg/cu.m was used to represent the exposure for the synthesis of the studies described above. 

Using this TWA and taking occupational ventilation rates and workweek into account results in a

LOAEL(HEC) of 0.009 mg/cu.m. 

 I.B.3. UNCERTAINTY AND MODIFYING FACTORS (INHALATION RfC)

UF -- An uncertainty factor of 10 was used for the protection of sensitive 

human subpopulations (including concern for acrodynia - see Additional 

Comments section) together with the use of a LOAEL. An uncertainty factor of 

3 was used for lack of data base, particularly developmental and reproductive 

studies. 

MF -- None

 I.B.4. ADDITIONAL COMMENTS (INHALATION RfC)

Probably the most widely recognized form of hypersensitivity to mercury poisoning is the

uncommon syndrome known as acrodynia, also called erythredema polyneuropathy or pink disease

(Warkany and Hubbard, 1953). Infantile acrodynia was first described in 1828, but adult cases have also

since been reported. While acrodynia has generally been associated with short-term exposures and with

urine levels of 50 ug/L or more, there are some cases in the literature in which mercury exposure was

known to have occurred, but no significant (above background) levels in urine were reported. There

could be many reasons for this, but the most likely is that urine levels are not a simple measure of body

burden or of target tissue (i.e., brain levels); however, they are the best means available for assessing the

extent of exposure. It was felt that the RfC level estimated for mercury vapor based on neurotoxicity of

chronic exposure in workers is adequate to protect children from risk of acrodynia because such

exposures of long duration would be expected to raise urine levels by only 0.12 ug/L against a

background level of up to 20 ug/L (i.e., such exposures would not add significantly to the background

level of mercury in those exposed).

Roels et al. (1987) investigated the relationships between the concentrations of metallic mercury

in air and levels monitored in blood or urine in workers exposed during manufacturing of dry alkaline

batteries. Breathing zone personal samples were used to characterize airborne mercury vapors. Total

mercury in blood and urine samples were analyzed using atomic absorption. The investigation controlled
B-5

for several key factors including the use of reliable personal air monitoring, quality control for blood and

urine analyses, standardization of the urinary mercury concentration for creatinine concentration, and

stability of exposure conditions (examined subjects were exposed to mercury vapor for at least 1 year). 

Strong correlations were found between the daily intensity of exposure to mercury vapor and the end of

workshift levels in blood (r=0.86; n=40) or urine (r=0.81; n=34). These relationships indicated a

conversion factor of 1:4.5 (air:blood) and 1:1.22 (air:urine as ug/g creatinine). These factors were used

to extrapolate blood or urine levels associated with effects in the reported studies to airborne mercury

levels.

Sensory and motor nerve conduction velocities were studied in 18 workers from a mercury cell

chlorine plant (Levine et al., 1982). Time-integrated urine Hg levels were used as an indicator of

mercury exposure. Using linearized regression analysis, the authors found that motor and sensory nerve

conduction velocity changes (i.e., prolonged distal latencies correlated with the time-integrated urinary

Hg levels in asymptomatic exposed workers) occurred when urinary Hg levels exceeded 25 ug/L. This

study demonstrates that mercury exposure can be associated with preclinical evidence of peripheral

neurotoxicity.

Singer et al. (1987) studied nerve conduction velocity of the median motor, median sensor and

sural nerves in 16 workers exposed to various inorganic mercury compounds (e.g., mercuric oxides,

mercurial chlorides, and phenyl mercuric acid) for an average of 7.3 +/- 7.1 years as compared with an

unexposed control group using t-tests. They found a slowing of nerve conduction velocity in motor, but

not sensory, nerves that correlated with increased blood and urine Hg levels and an increased number of

neurologic symptoms. The mean mercury levels in the exposed workers were 1.4 and 10 ug/L for blood

and urine, respectively. These urine levels are 2-fold less than those associated with peripheral

neurotoxicity in other studies (e.g., Levine et al., 1982). There was considerable variability in the data

presented by Singer et al. (1987), however, and the statistical analyses (t-test) were not as rigorous as

those employed by Levine et al. (1982) (linearized regression analysis). Furthermore, the subjects in the

Levine et al. (1982) study were asymptomatic at higher urinary levels than those reported to be

associated with subjective neurological complaints in the workers studied by Singer et al. (1987). 

Therefore, these results are not considered to be as reliable as those reported by Levine et al. (1982).

Miller et al. (1975) investigated several subclinical parameters of neurological dysfunction in

142 workers exposed to inorganic mercury in either the chloralkali industry or a factory for the

manufacture of magnetic materials. They reported a significant increase in average forearm tremor

frequency in workers whose urinary Hg concentrations exceeded 50 ug/L as compared with unexposed

controls. Also observed were eyelid fasciculation, hyperactive deep-tendon reflexes and dermatographia,

but there was no correlation between the incidence of these findings and urinary Hg levels.

Roels et al. (1985) examined 131 male and 54 female workers occupationally exposed to

mercury vapor for an average duration of 4.8 years. Urinary mercury (52 and 37 ug/g creatinine for

males and females, respectively) and blood mercury levels (14 and 9 ug/L for males and females,

respectively) were recorded, but atmospheric mercury concentration was not provided. Symptoms

indicative of CNS disorders were reported but not related to mercury exposure. Minor renal tubular

effects were detected in mercury-exposed males and females and attributed to current exposure intensity

rather (urinary Hg >50 ug/g creatinine) than exposure duration. Male subjects with urinary mercury

levels of >50 ug/g creatinine exhibited preclinical signs of hand tremor. It was noted that females did not

exhibit this effect and that their urinary mercury never reached the level of 50 ug/g creatinine. A

companion study (Roels et al., 1987) related air mercury (Hg-air)levels to blood mercury (Hg-blood) and
B-6

urinary mercury (Hg-U) values in 10 workers in a chloralkali battery plant. Duration of exposure was not

specified. A high correlation was reported for Hg-air and Hg-U for preshift exposure (r=0.70, p<0.001)

and post-shift (r=0.81, p<0.001) measurements. Based on these data and the results of their earlier

(1985) study, the investigators suggested that some mercury-induced effects may occur when Hg-U levels

exceed 50 ug/g creatinine, and that this value corresponds to a mercury TWA of about 40 ug/cu.m.

A survey of 567 workers at 21 chloralkali plants was conducted to ascertain the effects of

mercury vapor inhalation (Smith et al., 1970). Mercury levels ranged from <0.01 to 0.27 mg/cu.m and

chlorine concentrations ranged from 0.1 to 0.3 ppm at most of the working stations of these plants. 

Worker exposure to mercury levels (TWA) varied, with 10.2% of the workers being exposed to <0.01

mg/cu.m, 48.7% exposed to 0.01 to 0.05 mg/cu.m, 25.6% exposed to 0.06 to 0.10 mg/cu.m and 4.8%

exposed to 0.24 to 0.27 mg/cu.m (approximately 85% were exposed to Hg levels less than or equal to 0.1

mg/cu.m). The duration of employment for the examined workers ranged from one year (13.3%) to >10

years (31%), with 55.7% of the workers being employed for 2 or 9 years. A group of 600 workers not

exposed to chlorine served as a control group for assessment of chlorine effects, and a group of 382

workers not exposed to either chlorine or mercury vapor served as the reference control group. A strong

positive correlation (p<0.001) was found between the mercury TWAs and the reporting of subjective

neuropsychiatric symptoms (nervousness, insomnia), occurrence of objective tremors, and weight and

appetite loss. A positive correlation (p<0.001) was also found between mercury exposure levels and

urinary and blood mercury levels of test subjects. No adverse alterations in cardiorespiratory,

gastrointestinal, renal or hepatic functions were attributed to the mercury vapor exposure. Additionally,

biochemical (hematologic data, enzyme activities) and clinical measurements (EKG, chest X-rays) were

no different between the mercury-exposed and non-exposed workers. No significant signs or symptoms

were noted for individuals exposed to mercury vapor concentrations less than or equal to 0.1 mg/cu.m. 

This study provides data indicative of a NOAEL of 0.1 mg Hg/cu.m and a LOAEL of 0.18 mg Hg/cu.m. 

In a followup study conducted by Bunn et al. (1986), however, no significant differences in the frequency

of objective or subjective findings such as weight loss and appetite loss were observed in workers

exposed to mercury at levels that ranged between 50 and 100 ug/L. The study by Bunn et al. (1986) was

limited, however, by the lack of information provided regarding several methodological questions such as

quality assurance measures and control of possible confounding variables.

The mercury levels reported to be associated with preclinical and symptomatic neurological

dysfunction are generally lower than those found to affect kidney function, as discussed below.

Piikivi and Ruokonen (1989) found no evidence of glomerular or tubular damage in 60

chloralkali workers exposed to mercury vapor for an average of 13.7 +/- 5.5 years as compared with their

matched referent controls. Renal function was assessed by measuring urinary albumin and

N-acetyl-beta-glucosaminidase (NAG) activity. The mean blood Hg level in the exposed workers was 14

ug/L and the mean urinary level was 17 ug/L. The authors extrapolated the NOAEL for kidney effects

based on these results of 0.025 mg/cu.m from blood levels using the conversion factor calculated by

Roels et al. (1987).

Stewart et al. (1977) studied urinary protein excretion in 21 laboratory workers exposed to 10-50

ug/cu.m of mercury. Their urinary level of mercury was about 35 ug/L. Increased proteinuria was found

in the exposed workers as compared with unexposed controls. When preventive measure were instituted

to limit exposure to mercury, proteinuria was no longer observed in the exposed technicians. 
B-7

Lauwerys et al. (1983) found no change in several indices of renal function (e.g., proteinuria,

albuminuria, urinary excretion of retinol-binding protein, aminoaciduria, creatinine in serum,

beta-2-microglobulin in serum) in 62 workers exposed to mercury vapor for an average of 5.5 years. The

mean urinary Hg excretion in the exposed workers was 56 ug/g creatinine, which corresponds to an

exposure level of about 46 ug/cu.m according to a conversion factor of 1:1.22 (air:urine [ug/g creatinine])

(Roels et al., 1987). Despite the lack of observed renal effects, 8 workers were found to have an

increased in serum anti-laminin antibodies, which can be indicative of immunological effects. In a

followup study conducted by Bernard et al. (1987), however, there was no evidence of increased serum

anti-laminin antibodies in 58 workers exposed to mercury vapor for an average of 7.9 years. These

workers had a mean urinary Hg excretion of 72 ug/g creatinine, which corresponds to an exposure levels

of about 0.059 mg/cu.m.

Stonard et al. (1983) studied renal function in 100 chloralkali workers exposed to inorganic

mercury vapor for an average of 8 years. No changes in the following urinary parameters of renal

function were observed at mean urinary Hg excretion rates of 67 ug/g creatinine: total protein, albumin,

alpha-1-acid glycoprotein, beta-2-microglobulin, NAG, and gamma-glutamyl transferase. When urinary

Hg excretion exceeded 100 ug/g creatinine, a small increase in the prevalence of higher activities of

NAG and gamma-glutamyl transferase was observed. 

The mercury levels reported to be associated with preclinical and symptomatic neurological

dysfunction and kidney effects are lower than those found to pulmonary function, as discussed below.

McFarland and Reigel (1978) described the cases of 6 workers who were acutely exposed (4-8

hours) to calculated metallic mercury vapor levels of 1.1 to 44 mg/cu.m. These men exhibited a

combination of chest pains, dyspnea, cough, hemoptysis, impairment of pulmonary function (reduced

vital capacity), diffuse pulmonary infiltrates and evidence of interstitial pneumonitis. Although the

respiratory symptoms resolved, all six cases exhibited chronic neurological dysfunction, presumably as a

result of the acute, high-level exposure to mercury vapor. 

Lilis et al. (1985) described the case of a 31-year-old male who was acutely exposed to high

levels of mercury vapor in a gold-extracting facility. Upon admission to the hospital, the patient

exhibited dyspnea, chest pain with deep inspiration, irregular infiltrates in the lungs and reduced

pulmonary function (forced vital capacity [FVC]). The level of mercury to which he was exposed is not

known, but a 24-hour urine collection contained 1900 ug Hg/L. Although the patient improved gradually

over the next several days, 11 months after exposure he still showed signs of pulmonary function

abnormalities (e.g., restriction and diffusion impairment).

Levin et al. (1988) described four cases of acute high-level mercury exposure during gold ore

purification. The respiratory symptoms observed in these four cases ranged from minimal shortness of

breath and cough to severe hypoxemia. The most severely affected patient exhibited mild interstitial lung

disease both radiographically and on pulmonary function testing. One patient had a urinary Hg level of

245 ug/L upon hospital admission. The occurrence of long-term respiratory effects in these patients

could not be evaluated since all but one refused follow-up treatment.

Ashe et al. (1953) reported that there was no histopathological evidence of respiratory damage in

24 rats exposed to 0.1 mg Hg/cu.m 7 hr/day, 5 days/week for 72 weeks. This is equivalent to a

NOAEL[HEC] of 0.07 mg/cu.m.
B-8

Kishi et al. (1978) observed no histopathological changes in the lungs of rats exposed to 3

mg/cu.m of mercury vapor 3 hours/day, 5 days/week for 12-42 weeks.

Beliles et al. (1967) observed no histopathological changes in the lungs of pigeons exposed to 0.1

mg/cu.m of mercury vapor 6 hours/day, 5 days/week for 20 weeks.

Neurological signs and symptoms (i.e., tremors) were observed in 79 workers exposed to metallic

mercury vapor whose urinary mercury levels exceeded 500 ug/L. Short-term memory deficits were

reported in workers whose urine levels were less than 500 ug/L (Langolf et al., 1978).

Impaired performance in mechanical and visual memory tasks and psychomotor ability tests was

reported by Forzi et al. (1978) in exposed workers whose urinary Hg levels exceeded 100 ug/L. 

Decreased strength, decreased coordination, increased tremor, decreased sensation and increased

prevalence of Babinski and snout reflexes were exhibited by 247 exposed workers whose urinary Hg

levels exceeded 600 ug/L. Evidence of clinical neuropathy was observed at urinary Hg levels that

exceeded 850 ug/L (Albers et al., 1988).

Preclinical psychomotor dysfunction was reported to occur at a higher incidence in 43 exposed

workers (mean exposure duration of 5 years) whose mean urinary excretion of Hg was 50 ug/L. Workers

in the same study whose mean urinary Hg excretion was 71 ug/L had a higher incidence of total

proteinuria and albuminuria (Roels et al., 1982). 

Postural and intention tremor was observed in 54 exposed workers (mean exposure duration of

7.7 years) whose mean urinary excretion of Hg was 63 ug/L (Roels et al., 1989). 

Verbeck et al. (1986) observed an increase in tremor parameters with increasing urinary

excretion of mercury in 21 workers exposed to mercury vapor for 0.5-19 years. The LOAEL for this

effect was a mean urinary excretion of 35 ug/g creatinine.

Rosenman et al. (1986) evaluated routine clinical parameters (physical exams, blood chemistry,

urinalysis), neuropsychological disorders, urinary NAG, motor nerve conduction velocities and

occurrence of lenticular opacities in 42 workers of a chemical plant producing mercury compounds. A

positive correlation (p<0.05 to p<0.001) was noted between urinary mercury (levels ranged from 100-250

ug/L) and the number of neuropsychological symptoms, and NAG excretions and the decrease in motor

nerve conduction velocities.

Evidence of renal dysfunction (e.g., increased plasma and urinary concentrations of

beta-galactosidase, increased urinary excretion of high-molecular weight proteins and a slightly increased

plasma beta-2-microglobulin concentration) was observed in 63 chloralkali workers. The incidence of

these effects increased in workers whose urinary Hg excretion exceeded 50 ug/g 

creatinine (Buchet et al., 1980).

Increased urinary NAG levels were found in workers whose urinary Hg levels exceeded 50 ug/L

(Langworth et al., 1992).

An increase in the concentration of urinary brush border proteins (BB-50) was observed in 20

workers whose mean urinary Hg excretion exceeded 50 ug/g creatinine (Mutti et al., 1985). 
B-9

Foa et al. (1976) found that 15 out of 81 chloralkali workers exposed to 60-300 ug/cu.m mercury

exhibited proteinuria.

An increased excretion of beta-glutamyl transpeptidase, indicative of renal dysfunction, was

found in 509 infants dermally exposed to phenylmercury via contaminated diapers (Gotelli et al., 1985).

Berlin et al. (1969) exposed rats, rabbits and monkeys to 1 mg/cu.m of mercury vapor for 4 hours

and measured the uptake and distribution of mercury in the brain as compared with animals injected

intravenously with the same doses of mercury as mercuric salts. Mercury accumulated in the brain

following inhalation exposure to metallic mercury vapor at levels that were 10 times higher than those

observed following intravenous injection of the same dose of mercury as mercuric salts. These results

demonstrate that mercury is taken up by the brain following inhalation of the vapor at higher levels than

other forms of mercury and that this occurs in all species studied.

Limited animal studies concerning inhalation exposure to inorganic mercury are available. The

results of a study conducted by Baranski and Szymczyk (1973) were reported in an English abstract. 

Adult female rats were exposed to metallic mercury vapor at 2.5 mg/cu.m for 3 weeks prior to

fertilization and during gestation days 7-20. A decrease in the number of living fetuses was observed in

the dams compared with unexposed controls, and all pups born to the exposed dams died by the sixth day

after birth. However, no difference in the occurrence of developmental abnormalities was observed

between exposed and control groups. The cause of death of the pups in the mercury-exposed group was

unknown, although an unspecified percentage of the deaths was attributed by the authors to a failure of

lactation in the dams. Death of pups was also observed in another experiment where dams were only

exposed prior to fertilization (to 2.5 mg/cu.m), which supports the conclusion that the high mortality in

the first experiment was due at least in part to poor health of the mothers. Without further information,

this study must be considered inconclusive regarding developmental effects.

The only other study addressing the developmental toxicology of mercury is the one reported in

abstract form by Steffek et al. (1987) and, as such, is included as a supporting study. Sprague-Dawley

rats (number not specified) were exposed by inhalation to mercury vapor at concentrations of 0.1, 0.5 or

1.0 mg/cu.m throughout the period of gestation (days 1-20) or during the period of organogenesis (days

10-15). The authors indicated the exposure protocols to be chronic and acute exposure, respectively. At

either exposure protocol, the lowest mercury level produced no detectable adverse effect. At 0.5

mg/cu.m, an increase in the number of resorptions (5/41) was noted for the acute group, and two of 115

fetuses exhibited gross cranial defects in the chronic group. At 1.0 mg/cu.m, the number of resorptions

was increased in acute (7/71) and chronic (19/38) groups and a decrease in maternal and fetal weights

also was detected in the chronic exposure group. No statistical analysis for these data was provided. A

LOAEL of 0.5 mg/cu.m is provided based on these data. 

Mishinova et al. (1980) investigated the course of pregnancy and parturition in 349 women

exposed via inhalation to metallic mercury vapors (unspecified concentrations) in the workplace as

compared to 215 unexposed women. The authors concluded that the rates of pregnancy and labor

complication were high among women exposed to mercury and that the effects depended on "the length

of service and concentration of mercury vapors." Lack of sufficient details preclude the evaluation of

dose-response relationships.

In a questionnaire that assessed the fertility of male workers exposed to mercury vapor,

Lauwerys et al. (1985) found no statistically significant change in the observed number of children born
B-10

to the exposed group compared with a matched control group. The urinary excretion of mercury in the

exposed workers ranged from 5.1 to 272.1 ug/g creatinine.

Another study found that exposure to metallic mercury vapor caused prolongation of estrus

cycles in animals. Baranski and Szymczyk (1973) reported that female rats exposed via inhalation to

mercury vapor at an average of 2.5 mg/cu.m, 6 hours/day, 5 days/week for 21 days experienced longer

estrus cycles than unexposed animals. In addition, estrus cycles during mercury exposure were longer

than normal estrus cycles in the same animals prior to exposure. Although the initial phase of the cycle

was protracted, complete inhibition of the cycle did not occur. During the second and third weeks of

exposure, these rats developed signs of mercury poisoning including restlessness, seizures and trembling

of the entire body. The authors speculated that the effects on the estrus cycle were caused by the action

of mercury on the CNS (i.e., damage to the hypothalamic regions involved in the control of estrus

cycling).

Renal toxicity has been reported following oral exposure to inorganic mercury salts in animals,

with the Brown-Norway rat appearing to be uniquely sensitive to this effect. These mercury-induced

renal effects in the Brown-Norway rat are the basis for the oral RfD for mercurial mercury. Several

investigators have produced autoimmune glomerulonephritis by administering HgCl2 to Brown-Norway

rats (Druet et al., 1978).

The current OSHA standard for mercury vapor is 0.05 mg/cu.m. NIOSH recommends a TWA

Threshold Limit Value of 0.05 mg/cu.m for mercury vapor. 

 I.B.5. CONFIDENCE IN THE INHALATION RfC

Study -- Medium

Data Base -- Medium

RfC -- Medium

Due to the use of a sufficient number of human subjects, the inclusion of appropriate control

groups, the exposure duration, the significance level of the reported results and the fact that exposure

levels in a number of the studies had to be extrapolated from blood mercury levels, confidence in the key

studies is medium. The LOAEL values derived from these studies can be corroborated by other human

epidemiologic studies. The adverse effects reported in these studies are in accord with the

well-documented effects of mercury poisoning. The lack of human or multispecies

reproductive/developmental studies precludes assigning a high confidence rating to the data base and

inadequate quantification of exposure levels. Based on these considerations, the RfC for mercury is

assigned a confidence rating of medium.

 I.B.6. EPA DOCUMENTATION AND REVIEW OF THE INHALATION RfC

Source Document -- U.S. EPA, 1995

This IRIS summary is included in The Mercury Study Report to Congress which was reviewed

by OHEA and EPA's Mercury Work Group in November 1994. An interagency review by scientists from
B-11

other federal agencies took place in January 1995. The report was also reviewed by a panel of

non-federal external scientists in January 1995 who met in a public meeting on January 25-26. All

reviewers comments have been carefully evaluated and considered in the revision and finalization of this

IRIS summary. A record of these comments is summarized in the IRIS documentation files.

Other EPA Documentation -- None

Agency Work Group Review -- 11/16/89, 03/22/90, 04/19/90

Verification Date -- 04/19/90

 I.B.7. EPA CONTACTS (INHALATION RfC)

Annie M. Jarabek / NCEA -- (919)541-4847

William F. Sette / OPP -- (703)305-6375

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occupational elemental mercury exposure. Ann. Neurol. 24(5): 651-659.

Ashe, W.F., E.J. Largent, F.R. Dutra, D.M. Hubbard and M. Blackstone. 1953. Behavior of mercury in

the animal organism following inhalation. Ind. Hyg. Occup. Med. 17: 19-43.

Baranski, B. and I. Szymczyk. 1973. [Effects of mercury vapor upon reproductive functions of female

white rats]. Med. Pr. 24(3): 249-261. (Czechoslovakian)

Beliles, R.P., R.S. Clark, P.R. Belluscio, C.L. Yuile and L.J. Leach. 1967. Behavioral effects in pigeons

exposed to mercury vapor at a concentration of 0.1 mg/cu.m. Am. Ind. Hyg. J. 28(5): 482-484.

Berlin, M., J. Fazackerley and G. Nordberg. 1969. The uptake of mercury in the brains of mammals

exposed to mercury vapor and to mercuric salts. Arch. Environ. Health. 18: 719-729. 

Bernard, A.M., H.R. Roels, J.M. Foldart and R.L. Lauwerys. 1987. Search for anti-laminin antibodies in

the serum of workers exposed to cadmium, mercury vapour or lead. Int. Arch. Occup. Environ. Health. 

59: 303-309.

Buchet, J.P., H. Roels, A. Bernard and R. Lauwerys 1980. Assessment of renal function of workers

exposed to inorganic lead, cadmium or mercury vapor. J. Occup. Med. 22(11): 741-750.

Bunn, W.B., C.M. McGill, T.E. Barber, J.W. Cromer and L.J. Goldwater. 1986. Mercury exposure in

chloralkali plants. Am. Ind. Hyg. Assoc. J. 47(5): 249-254. 

Druet, P., E. Druet, F. Potdevin, et al. 1978. Immune type glomerulonephritis induced by HgCl2 in the

Brown-Norway rat. Ann. Immunol. 129C: 777-792. 
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Fawer, R.F., Y. DeRibaupierre, M.P. Guillemin, M. Berode and M. Lob. 1983. Measurement of hand

tremor induced by industrial exposure to metallic mercury. J. Ind. Med. 40: 204-208.

Foa, V., L. Caimi, L. Amante, et al. 1976. Patterns of some lysosomal enzymes in the plasma and of

proteins in urine of workers exposed to inorganic mercury. Int. Arch. Occup. Environ. Health. 37:

115-124. 

Forzi, M., M.G. Cassitto, C. Bulgheroni and V. Foa. 1978. Psychological measures in workers

occupationally exposed to mercury vapors: A validation study. In: Adverse Effects of Environmental

Chemicals and Psychotropic Drugs: Neurophysiological and Behavioral Tests, Vol. 2, H.J. Zimmerman,

Ed. Appleton-Century-Crofts, New York, NY. p. 165-171.

Gotelli, C.A., E. Astolfi, C. Cox, E. Cernichiari and T. Clarkson. 1985. Early biochemical effects of an

organic mercury funcigicide on infants: "Dose makes the poison". Science. 277: 638-640. 

Kishi, R., K. Hashimoto, S. Shimizu and M. Kobayashi. 1978. Behavioral changes and mercury

concentrations in tissues of rats exposed to mercury vapor. Toxicol. Appl. Pharmacol. 46(3): 555-566.

Langolf, G.D., D.B. Chaffin, R. Henderson and H.P. Whittle. 1978. Evaluation of workers exposed to

elemental mercury using quantitative tests of tremor and neuromuscular functions. Am. Ind. Hyg. Assoc.

J. 39: 976-984.

Langworth, S., C.G. Elinder, K.G. Sundquist and O. Vesterberg. 1992. Renal and immunological effects

of occupational exposure to inorganic mercury. Br. J. Ind. Med. 49: 394-401.

Lauwerys, R., A. Bernard, H. Roels, et al. 1983. Anti-laminin antibodies in workers exposed to mercury

vapour. Toxicol. Lett. 17: 113-116. 

Lauwerys, R., H. Roels, P. Genet, G. Toussaint, A. Bouckaert and S. De Cooman. 1985. Fertility of

male workers exposed to mercury vapor or to manganese dust: A questionnaire study. Am. J. Ind. Med. 

7(2): 171-176.

Levin, M., J. Jacobs and P.G. Polos. 1988. Acute mercury poisoning and mercurial pneumonitis from

gold ore purification. Chest. 94(3): 554-558.

Levine, S.P., G.D. Cavender, G.D. Langolf and J.W. Albers. 1982. Elemental mercury exposure: 

Peripheral neurotoxicity. Br. J. Ind. Med. 39: 136-139. 

Liang, Y-X., R-K. Sun, Y. Sun, Z-Q. Chen and L-H. Li. 1993. Psychological effects of low exposure to

mercury vapor: Application of a computer-administered neurobehavioral evaluation system. Environ.

Res. 60: 320-327. 

Lilis, R., A. Miller and Y. Lerman. 1985. Acute mercury poisoning with severe chronic pulmonary

manifestations. Chest. 88(2): 306-309.

McFarland, R.B. and H. Reigel. 1978. Chronic mercury poisoning from a single brief exposure. J.

Occup. Med. 20(8): 532-534.
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Miller, J.M., D.B. Chaffin and R.G. Smith. 1975. Subclinical psychomotor and neuromuscular changes

in workers exposed to inorganic mercury. Am. Ind. Hyg. Assoc. J. 36: 725-733.

Mishonova, V.N., P.A. Stepanova and V.V. Zarudin. 1980. Characteristics of the course of pregnancy

and labor in women coming in contact with low concentrations of metallic mercury vapors in

manufacturing work places. Gig Tr Prof Zabol. Issue 2: 21-23. 

Mutti, A., S. Lucertini, M. Fornari, et al. 1985. Urinary excretion of a brush-border antigen revealed by

monoclonal antibodies in subjects occupationally exposed to heavy metals. Heavy Met Environ.

International Conference 5th. Vol.1. p. 565-567. 

Ngim, C.H., S.C. Foo, K.W. Boey and J. Jeyaratnam. 1992. Chronic neurobehavioral effects of

elemental mercury in dentists. Br. J. Ind. Med. 49: 782-790.

Piikivi, L. 1989. Cardiovascular reflexes and low long-term exposure to mercury vapor. Int. Arch.

Occup. Environ. Health. 61: 391-395.

Piikivi, L. and H. Hanninen. 1989. Subjective symptoms and psychological performance of

chlorine-alkali workers. Scand. J. Work Environ. Health. 15: 69-74.

Piikivi, L. and A. Ruokonen. 1989. Renal function and long-term low mercuryvapor exposure. Arch.

Environ. Health. 44(3): 146-149. 

Piikivi, L. and U. Tolonen. 1989. EEG findings in chlor-alkali workers subjected to low long term

exposure to mercury vapor. Br. J. Ind. Med. 46: 370-375.

Roels, H., R. Lauwerys, J.P. Buchet, et al. 1982. Comparison of renal function and psychomotor

performance in workers exposed to elemental mercury. Int. Arch. Occup. Environ. Health. 50: 77-93.

Roels, H., J.P. Gennart, R. Lauwreys, J.P. Buchet, J. Malchaire and A. Bernard. 1985. Surveillance of

workers exposed to mercury vapor: validation of a previously proposed biological threshold limit value

for mercury concentration in urine. Am. J. Ind. Med. 7: 45-71. 

Roels, H., S. Abdeladim, E. Ceulemans and R. Lauwreys. 1987. Relationships between the

concentrations of mercury in air and in blood or urine in workers exposed to mercury vapour. Ann.

Occup. Hyg. 31(2): 135-145.

Roels, H., S. Abdeladim, M. Braun, J. Malchaire and R. Lauwerys. 1989. Detection of hand tremor in

workers exposed to mercury vapor: A comparative study of three methods. Environ. Res. 49: 152-165.

Rosenman, K.D., J.A. Valciukas, L. Glickman, B.R. Meyers and A. Cinotti. 1986. Sensitive indicators

of inorganic mercury toxicity. Arch. Environ. Health. 41(4): 208-215.

Singer, R., J.A. Valciukas and K.D. Rosenman. 1987. Peripheral neurotoxicity in workers exposed to

inorganic mercury compounds. Arch. Environ. Health. 42(4): 181-184.

Smith, R.G., A.J. Vorwald, L.S. Patil and T.F. Mooney, Jr. 1970. Effects of exposure to mercury in the

manufacture of chlorine. Am. Ind. Hyg. Assoc. J. 31(1): 687-700.
B-14

Steffek, A.J., R. Clayton, C. Siew and A.C. Verrusio. 1987. Effects of elemental mercury vapor

exposure on pregnant Sprague-Dawley rats (abstract only). Teratology. 35: 59A.

Stewart, W.K., H.A. Guirgis, J. Sanderson and W. Taylor. 1977. Urinary mercury excretion and

proteinuria in pathology laboratory staff. Br. J. Ind. Med. 34: 26-31.

Stonard, M.D., B.V. Chater, D.P. Duffield, A.L. Nevitt, J.J. O'Sullivan and G.T. Steel. 1983. An

evaluation of renal function in workers occupationally exposed to mercury vapor. Int. Arch. Occup.

Environ. Health. 52: 177-189.

U.S. EPA. 1995. Mercury Study Report to Congress. Office of Research and Development, Washington

DC 20460. EPA/600/P-94/002Ab. External Review Draft. 

Verbeck, M.M., H.J.A. Salle and C.H. Kemper. 1986. Tremor in workers with low exposure to metallic

mercury. Hyg. Assoc. J. 47(8): 559-562. 

Warkany, J. and D.M. Hubbard. 1953. Acrodynia and mercury. J. Pediat. 42:365-386.
B-15

 II. CARCINOGENICITY ASSESSMENT FOR LIFETIME EXPOSURE

Substance Name -- Mercury, elemental

CASRN -- 7439-97-6

Preparation Date -- 5/24/94

 II.A. EVIDENCE FOR CLASSIFICATION AS TO HUMAN CARCINOGENICITY

 II.A.1 WEIGHT-OF-EVIDENCE CLASSIFICATION

Classification -- D; not classifiable as to human carcinogenicity

Basis -- Based on inadequate human and animal data. Epidemiologic studies failed to show a correlation

between exposure to elemental mercury vapor and carcinogenicity; the findings in these studies were

confounded by possible or known concurrent exposures to other chemicals, including human

carcinogens, as well as lifestyle factors (e.g., smoking). Findings from genotoxicity tests are severely

limited and provide equivocal evidence that mercury adversely affects the number or structure of

chromosomes in human somatic cells. 

 II.A.2 HUMAN CARCINOGENICITY DATA 

Inadequate. A number of epidemiological studies were conducted that examined mortality

among elemental mercury vapor-exposed workers. Conflicting data regarding a correlation between

mercury exposure and an increased incidence of cancer mortalities have been obtained. All of the studies

have limitations that complicate interpretation of their results for associations between mercury exposure

and induction of cancer; increased cancer rates were attributable to other concurrent exposures or

lifestyle factors.

A retrospective cohort study examined mortality among 5663 white males who worked between

1953 and 1963 at a plant in Oak Ridge, Tennessee, where elemental mercury was used for lithium isotope

separation (Cragle et al., 1984). The workers were divided into three cohorts: exposed workers who had

been monitored on a quarterly basis for mercury levels in urine (n=2,133); workers exposed in the

mercury process section for whom urinalysis monitoring data were not collected (n=270); and unexposed

workers from other sections of the nuclear weapons production facility (n=3260). The study subjects

worked at least 4 months during 1953-1958 (a period when mercury exposures were likely to be high);

mortality data from death certificates were followed through the end of 1978. The mean age of the men

at first employment at the 

facility was 33 years, and the average length of their employment was >16 years with a mean of 3.73

years of estimated mercury exposure. Air mercury levels were monitored beginning in 1955; during

1955 through the third quarter of 1956, air mercury levels were reportedly above 100 ug/cu.m in 30-80%

of the samples. Thereafter, air mercury levels decreased to concentrations below 100 ug/cu.m. The

mortality experience (i.e., the SMR) of each group was compared with the age-adjusted mortality

experience of the U.S. white male population. Among exposed and monitored workers, no significant

increases in mortality from cancer at any site were reported, even after the level or length of exposure
B-16

was considered. A significantly lower mortality from all causes was observed. An excessive number of

deaths was reportedly due to lung cancer in the exposed and monitored workers (42 observed, 31.36

expected), but also in the unexposed workers (71 observed, 52.93 expected). The SMR for each group

was 1.34; the elevated incidence of lung cancer deaths was, therefore, attributed to some other factor at

the plant and/or to lifestyle factors (e.g., smoking) common to both the exposed and unexposed groups. 

Study limitations include small cohort sizes for cancer mortality, which limited the statistical stability of

many comparisons.

Barregard et al. (1990) studied mortality and cancer morbidity between 1958 and 1984 in 1190

workers from eight Swedish chloralkali plants that used the mercury cell process in the production of

chlorine. The men included in the study had been monitored for urinary or blood mercury for more than

one year between 1946 and 1984. Vital status and cause of death were ascertained from the National

Population Register and the National Bureau of Statistics. The cancer incidence of the cohort was

obtained from the Swedish Cancer Register. The observed total mortality and cancer incidences were

compared with those of the general Swedish male population. Comparisons were not made between

exposed and unexposed workers. Mean urinary mercury levels indicated a decrease in exposure between

the 1950s and 1970s; the mean urinary mercury level was 200 ug/L during the 1950s, 150 ug/L during

the 1960s and 50 ug/L in the 1970s. Mortality from all causes was not significantly increased in exposed

workers. A significant increase in deaths from lung tumors was observed in exposed workers 10 years or

more after first exposure (rate ratio, 2.0; 95% CI, 1.0-3.8). Nine of the 10 observed cases of lung cancer

occurred among workers (457 of the 1190) possibly exposed to asbestos as well as to mercury. No dose

response was observed with respect to mercury exposure and lung tumors. This study is limited because

no quantitation was provided on smoking status, and results were confounded by exposure to asbestos.

Ahlbom et al. (1986) examined the cancer mortality during 1961-1979 of cohorts of Swedish

dentists and dental nurses aged 20-64 and employed in 1960 (3454 male dentists, 1125 female dentists,

4662 female dental nurses). Observed incidences were compared with those expected based on cancer

incidence during 1961-1979 among all Swedes employed during 1960 and the proportion of all Swedes

employed as dentists and dental nurses. Data were stratified by sex, age (5-year age groups) and county. 

The incidence of glioblastomas among the dentists and dental nurses combined was significantly

increased compared to survival rates (SMR, 2.1; 95% CI, 1.3-3.4); the individual groups had apparently

elevated SMRs (2.0-2.5), but the 95% 

confidence intervals of these groups included unity. By contrast, physicians and nurses had SMRs of

only 1.3 and 1.2, respectively. Exposure to mercury could not be established as the causative factor

because exposure to other chemicals and X-rays was not ruled out.

Amandus and Costello (1991) examined the association between silicosis and lung cancer

mortality between 1959 and 1975 in 9912 white male metal miners employed in the United States

between 1959 and 1961. Mercury exposures were not monitored. Exposures to specific metals among

the silicotic and nonsilicotic groups were analyzed separately. Lung cancer mortality in both silicotic

and nonsilicotic groups was compared with rates in white males in the U.S. population. Both silicotic

(n=11) and nonsilicotic mercury miners (n=263) had significantly increased lung cancer mortality (SMR,

14.03; 95% CI, 2.89-40.99 for silicotics. SMR, 2.66; 95% CI, 1.15-5.24 for nonsilicotics). The analysis

did not focus on mercury miners, and confounders such as smoking and radon exposure were not

analyzed with respect to mercury exposure. This study is also limited by the small sample size for

non-silicotic mercury miners.
B-17

A case-control study of persons admitted to a hospital in Florence, Italy, with lung cancer

between 1981-1983 was performed to evaluate occupational risk factors (Buiatti et al., 1985). Cases

were matched with one or two controls (persons admitted to the hospital with diagnoses other than lung

cancer or suicide) with respect to sex, age, date of admission and smoking status. Women who had "ever

worked" as hat makers had a significantly increased risk of lung cancer. The duration of employment as

a hat maker averaged 22.2 years, and latency averaged 47.8 years. Workers in the Italian hat industry

were known to be occupationally exposed to mercury; however, the design of this study did not allow

evaluation of the relationship between cumulative exposure and cancer incidence. In addition,

interpretation of the results of this study is limited by the small sample size (only 6/376 cases reported

this occupation) and by exposure of hat makers to other pollutants including arsenic, a known lung

carcinogen.

Ellingsen et al. (1992) examined the total mortality and cancer incidence among 799 workers

employed for more than 1 year in two Norwegian chloralkali plants. Mortality incidence between 1953

and 1988 and cancer incidence between 1953 and 1989 were examined. Mortality and cancer incidence

were compared with that of the age-adjusted general male Norwegian population. No increase in total

cancer incidence was reported, but lung cancer was significantly elevated in the workers (rate ratio, 1.66;

95% CI, 1.0-2.6). No causal relationship can be drawn from the study between mercury exposure and

lung cancer because no correlation existed between cumulative mercury dose, years of employment or

latency time. Also, the prevalence of smoking was 10 20% higher in the exposed workers, and many

workers were also exposed to asbestos. 

 II.A.3 ANIMAL CARCINOGENICITY DATA

Inadequate. Druckrey et al. (1957) administered 0.1 mL of metallic mercury to 39 male and

female rats (BD III and BD IV strains) via intraperitoneal injection. Among the rats surviving longer

than 22 months, 5/12 developed peritoneal sarcomas. The increase in the incidence of sarcomas was

observed only in those tissues that had been in direct contact with the mercury. Although severe kidney

damage was reported in all treated animals, no renal tumors or tumors at any site other than the peritoneal

cavity were observed. 

 II.A.4 SUPPORTING DATA FOR CARCINOGENICITY

Cytogenetic monitoring studies of workers occupationally exposed to mercury by inhalation

provide very limited evidence that mercury adversely affects the number or structure of chromosomes in

human somatic cells. Popescu et al. (1979) compared four men exposed to elemental mercury vapor with

an unexposed group and found a statistically significant increase in the incidence of chromosome

aberrations in the WBCs from whole blood. Verschaeve et al. (1976) found an increase in aneuploidy

after exposure to low concentrations of vapor, but results could not be repeated in later studies

(Verschaeve et al., 1979). Mabille et al. (1984) did not find increases in structural chromosomal

aberrations of lymphocytes of exposed workers. Similarly, Barregard et al. (1991) found no increase in

the incidence or size of micronuclei and no correlation between micronuclei and blood or urinary

mercury levels of chloralkali workers. A statistically significant correlation was observed between
B-18

cumulative exposure to mercury and micronuclei induction in T lymphocytes in exposed workers,

suggesting a genotoxic effect. 

 II.B QUANTITATIVE ESTIMATE OF CARCINOGENIC RISK FROM ORAL EXPOSURE

None.

 II.C QUANTITATIVE ESTIMATE OF CARCINOGENIC RISK FROM INHALATION

EXPOSURE

None.

 II.D EPA DOCUMENTATION, REVIEW, AND CONTACTS (CARCINOGENICITY

ASSESSMENT)

 II.D.1 EPA DOCUMENTATION

Source document -- U.S. EPA, 1995

This IRIS summary is included in The Mercury Study Report to Congress which was reviewed

by OHEA and EPA's Mercury Work Group in November 1994. An interagency review by scientists from

other federal agencies took place in January 1995. The report was also reviewed by a panel of

non-federal external scientists in January 1995 who met in a public meeting on January 25-26. All

reviewers comments have been carefully evaluated and considered in the revision and finalization of this

IRIS summary. A record of these comments is 

summarized in the IRIS documentation files.

 II.D.2 REVIEW (CARCINOGENICITY ASSESSMENT)

Agency Work Group Review -- 01/13/88, 03/03/94

Verification Date -- 03/03/94

 II.D.3 U.S. EPA CONTACTS (CARCINOGENICITY ASSESSMENT)

Rita Schoeny / NCEA -- (513)569-7544

REFERENCES
B-19

Ahlbom, A., S. Norell, Y. Rodvall and M. Nylander. 1986. Dentists, dental nurses, and brain tumours. 

Br. Med. J. 292: 662. 

Amandus, H. and J. Costello. 1991. Silicosis and lung cancer in U.S. metal miners. Arch. Environ.

Health. 46(2): 82-89. 

Barregard, L., G. Sallsten and B. Jarvholm. 1990. Mortality and cancer incidence in chloralkali workers

exposed to inorganic mercury. Br. J. Ind. Med. 47(2): 99-104. 

Barregard, L., B. Hogstedt, A. Schutz, A. Karlsson, G. Sallsten and G. Thiringer. 1991. Effects of

occupational exposure to mercury vapor on lymphocyte micronuclei. Scand. J. Work Environ. Health. 

17: 263-268.

Buiatti, E., D. Kriebel, M. Geddes, M. Santucci and N. Pucci. 1985. A case control study of lung cancer

in Florence, Italy. I. Occupational risk factors. J. Epidemiol. Comm. Health. 39: 244-250. 

Cragle, D.L., D.R. Hollis, J.R. Qualters, W.G. Tankersley and S.A. Fry. 1984. A mortality study of men

exposed to elemental mercury. J. Occup. Med. 26(11): 817-821. 

Druckrey, H., H. Hamperl and D. Schmahl. 1957. Carcinogenic action of metallic mercury after

intraperitoneal administration in rats. Z. Krebsforsch. 61: 511-519. (Cited in U.S. EPA, 1985) 

Ellingsen, D., A. Andersen, H.P. Nordhagen, J. Efskind and H. Kjuus. 1992. Cancer incidence and

mortality among workers exposed to mercury in the Norwegian chloralkali industry. 8th International

Symposium on Epidemiology in Occupational Health, Paris, France, September 10-12, 1991. Rev.

Epidemiol. Sante Publique. 40(1): S93-S94. 

Mabille, V., H. Roels, P. Jacquet, A. Leonard and R. Lauwerys. 1984. Cytogenetic examination of

leucocytes of workers exposed to mercury vapor. Int. Arch. Occup. Environ. Health. 53: 257-260. 

Popescu, H.I., L. Negru and I. Lancranjan. 1979. Chromosome aberrations induced by occupational

exposure to mercury. Arch. Environ. Health. 34(6): 461-463. 

U.S. EPA. 1980. Ambient Water Quality Criteria Document for Mercury. Prepared by the Office of

Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH

for the Office of Water Regulation and Standards, Washington, DC. EPA/440/5-80/058. NTIS PB

81-117699. 

U.S. EPA. 1984a. Mercury Health Effects Update: Health Issue Assessment. Final Report. Prepared

by the Office of Health and Environmental Assessment, Environmental Criteria and Assessment Office,

Cincinnati, OH for the Office of Air Quality Planning and Standards, Research Triangle Park, NC. 

EPA/600/8-84/019F. NTIS PB81-85-123925. 

U.S. EPA. 1984b. Health Effects Assessment for Mercury. Prepared by the Office of Health and

Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH for the

Office of Emergency and Remedial Response, Washington, DC. EPA/540/1086/042. NTIS

PB86-134533/AS. 
B-20

U.S. EPA. 1985. Drinking Water Criteria Document for Mercury. Prepared by the Office of Health and

Environmental Assessment Office, Cincinnati, OH for the Office of Drinking Water, Washington, DC. 

EPA/600/X-84/178. NTIS PB86-117827. 

U.S. EPA. 1988. Drinking Water Criteria Document for Inorganic Mercury. Prepared by the Office of

Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH

for the Office of Drinking Water, Washington, DC. EPA/600/X-84/178. NTIS PB89-192207. 

U.S. EPA. 1993. Summary Review of Health Effects Associated with Mercuric Chloride: Health Issue

Assessment (Draft). Prepared by the Office of Health and Environmental Assessment, Environmental

Criteria and Assessment Office, Cincinnati, OH for the Office of Air Quality Planning and Standards,

Research Triangle Park, NC. EPA/600/R-92/199. 

U.S. EPA. 1995. Mercury Study Report to Congress. Office of Research and Development,

Washington, DC. External Review Draft. EPA/600/P-94/002Ab.

Verschaeve, L., M. Kirsch-Volders, C. Susanne et al. 1976. Genetic damage induced by occupationally

low mercury exposure. Environ. Res. 12: 303-316. 

Verschaeve, L., J.P. Tassignon, M. Lefevre, P. De Stoop and C. Susanne. 1979. Cytogenetic

investigation on leukocytes of workers exposed to metallic mercury. Environ. Mutagen. 1: 259-268. 
B-21

 I.A REFERENCE DOSE FOR CHRONIC ORAL EXPOSURE (RfD)

Substance Name -- Mercuric chloride (HgCl2)

CASRN -- 7487-94-7

Preparation Date -- 11/01/88

 I.A.1 ORAL RfD SUMMARY

Critical Effect Experimental Doses* UF MF RfD

Autoimmune effects NOAEL: None 1000 1 3E-4

mg/kg-day

Rat Subchronic LOAEL: 0.226 mg/kg-day

Feeding and

Subcutaneous LOAEL: 0.317 mg/kg-day

Studies

LOAEL: 0.633 mg/kg-day

U.S. EPA, 1987

* Conversion Factors and Assumptions -- Dose conversions in the three studies employed a 0.739 factor

for HgCl2 to Hg2+, a 100% factor for subcutaneous (s.c.) to oral route of exposure, and a time-weighted

average for days/week of dosing. This RfD is based on the back calculations from a Drinking Water

Equivalent Level (DWEL), recommended to and subsequently adopted by the Agency, of 0.010 mg/L:

(RfD = 0.010 mg/L x 2 L/day/70 kg bw = 0.0003 mg/kg bw/day). The LOAEL exposure levels, utilized

in the three studies selected as the basis of the recommended DWEL, are from Druet et al. (1978),

Bernaudin et al. (1981) and Andres (1984), respectively.

 I.A.2 PRINCIPAL AND SUPPORTING STUDIES (ORAL RfD)

U.S. EPA. 1987. Peer Review Workshop on Mercury Issues. Summary Report. Environmental Criteria

and Assessment Office, Cincinnati, OH. October 26-27. 

On October 26-27, 1987, a panel of mercury experts met at a Peer Review Workshop on Mercury

Issues in Cincinnati, Ohio, and reviewed outstanding issues concerning the health effects and risk

assessment of inorganic mercury (U.S. EPA, 1987). The following five consensus conclusions and

recommendations were agreed to as a result of this workshop:

1) The most sensitive adverse effect for mercury risk assessment is formation of

mercuric-mercury-induced autoimmune glomerulonephritis. The production and

deposition of IgG antibodies to the glomerular basement membrane can be considered

the first step in the formation of this mercuric-mercury-induced autoimmune

glomerulonephritis. 
B-22

2) The Brown Norway rat should be used for mercury risk assessment. The Brown Norway

rat is a good test species for the study of Hg2+-induced autoimmune glomerulonephritis. 

The Brown Norway rat is not unique in this regard (this effect has also been observed in

rabbits).

3) The Brown Norway rat is a good surrogate for the study of mercury-induced kidney

damage in sensitive humans. For this reason, the uncertainty factor used to calculate

criteria and health advisories (based on risk assessments using the Brown Norway rat)

should be reduced by 10-fold.

4) Hg2+ absorption values of 7% from the oral route and 100% from the s.c. route should

be used to calculate criteria and health advisories.

5) A DWEL of 0.010 mg/L was recommended based on the weight of evidence from the

studies using Brown Norway rats and limited human tissue data.

Three studies using the Brown Norway rat as the test strain were chosen from a larger selection of studies

as the basis for the panel's recommendation of 0.010 mg/L as the DWEL for inorganic mercury. The

three studies are presented below for the sake of completeness. It must be kept in mind, however, that

the recommended DWEL of 0.010 mg/L and back calculated oral RfD of 0.0003 mg/kg-day were arrived

at from an intensive review and workshop discussions of the entire inorganic mercury data base, not just

from one study.

In the Druet et al. (1978) study, the duration of exposure was 8-12 weeks; s.c. injection was used

instead of oral exposure. In this study the development of kidney disease was evaluated. In the first

phase the rats developed anti-GBM antibodies. During the second phase, which is observed after 2-3

months, the patterns of fixation of antisera changed from linear to granular as the disease progressed. 

The immune response was accompanied by proteinuria and in some cases by a nephrotic syndrome. 

Both male and female Brown Norway rats 7-9 weeks of age were divided into groups of 6-20

animals each. The numbers of each sex were not stated. The animals received s.c. injections of mercuric

chloride (HgCl2) 3 times weekly for 8 weeks, with doses of 0, 100, 250, 500, 1000 and 2000 ug/kg. An

additional group was injected with a 50 ug/kg dose for 12 weeks. Antibody formation was measured by

the use of kidney cryostat sections stained with a fluoresceinated sheep anti-rat IgG antiserum. Urinary

protein was assessed by the biuret method (Druet et al., 1978).

Tubular lesions were observed at the higher dose levels. Proteinuria was reported at doses of 100

ug/kg and above, but not at 50 ug/kg. Proteinuria was considered a highly deleterious effect, given that

affected animals developed hypoalbuminemia and many died. Fixation of IgG antiserum was detected in

all groups except controls (Druet et al., 1978).

Bernaudin et al. (1981) reported that mercurials administered by inhalation or ingestion to Brown

Norway rats developed a systemic autoimmune disease. The HgCl2 ingestion portion of the study

involved the forcible feeding of either 0 or 3000 ug/kg-week of HgCl2 to male and female Brown

Norway rats for up to 60 days. No abnormalities were reported using standard histological techniques in

either experimental or control rats. Immunofluorescence histology revealed that 80% (4/5) of the

mercuric-exposed rats were observed with a linear IgG deposition in the glomeruli after 15 days of

exposure. After 60 days of HgCl2 exposure, 100% (5/5) of the rats were seen with a mixed linear and
B-23

granular pattern of IgG deposition in the glomeruli and granular IgG deposition in the arteries. Weak

proteinuria was observed in 60% (3/5) of the rats fed HgCl2 for 60 days. The control rats were observed

to have no deposition of IgG in the glomeruli or arteries as well as normal urine protein concentrations. 

Andres (1984) administered HgCl2 (3 mg/kg in 1 mL of water) by gavage to five Brown Norway

rats and two Lewis rats twice a week for 60 days. A sixth Brown Norway rat was given only 1 mL of

water by gavage twice a week for 60 days. All rats had free access to tap water and pellet food. After

2-3 weeks of exposure, the Brown Norway HgCl2-treated rats started to lose weight and hair. Two of the

HgCl2-treated Brown Norway rats died 30-40 days after beginning the study. No rats were observed to

develop detectable proteinuria during the 60-day study. The kidneys appeared normal in all animals

when evaluated using standard histological techniques, but examination by immunofluorescence showed

deposits of IgG present in the renal glomeruli of only the mercuric-treated Brown Norway rats. The

Brown Norway treated rats were also observed with mercury-induced morphological lesions of the ileum

and colon with abnormal deposits of IgA in the basement membranes of the intestinal glands and of IgG

in the basement membranes of the lomina propria. All observations in the Lewis rats and the control

Brown Norway rat appeared normal. 

 I.A.3 UNCERTAINTY AND MODIFYING FACTORS (ORAL RfD)

UF -- An uncertainty factor of 1000 was applied to the animal studies using Brown Norway rats as

recommended in U.S. EPA (1987). An uncertainty factor was applied for LOAEL to NOAEL

conversion: 10 for use of subchronic studies and a combined 10 for both animal to human and sensitive

human populations.

MF -- None

 I.A.4 ADDITIONAL STUDIES / COMMENTS (ORAL RfD)

Kazantzis et al. (1962) performed renal biopsies in 2 (out of 4) workers with nephrotic syndrome

who had been occupationally exposed to mercuric oxide, mercuric acetate and probably mercury vapors. 

Investigators reported that the nephrotic syndrome observed in 3 of the 4 workers may have been an

idiosyncratic reaction since many other workers in a factory survey had similarly high levels of urine

mercury without developing proteinuria. This conclusion was strengthened by work in Brown Norway

rats indicating a genetic (strain) susceptibility and that similar mercury-induced immune system

responses have been seen in affected humans and the susceptible Brown Norway rats (U.S. EPA, 1987).

The only chronic ingestion study designed to evaluate the toxicity of mercury salts was reported

by Fitzhugh et al. (1950). In this study, rats of both sexes (20-24/group) were given 0.5, 2.5, 10, 40 or

160 ppm mercury as mercuric acetate in their food for up to 2 years. Assuming food consumption was

equal to 5% bw/day, the daily intake would have been 0.025, 0.125, 0.50, 2.0 and 8.0 mg/kg for the five

groups, respectively. At the highest dose level, a slight depression of body weight was detected in male

rats only. The statistical significance of this body-weight depression was not stated. Kidney weights

were significantly (p<0.05) increased at the 2.0 and 8.0 mg/kg dose levels. Pathological changes

originating in the proximal convoluted tubules of the kidneys were also noted, with more severe effects
B-24

in females than males. The primary weaknesses of this study were (1) the lack of reporting on which

adverse effects were observed with which dosing groups and (2) that the most sensitive strain, the Brown

Norway rat, was not used for evaluating the mercury-induced adverse health effects. 

NTP (1993) conducted subchronic and chronic gavage toxicity studies on Fischer 344 rats and

B6C3F1 mice to evaluate the effects of HgCl2, and the kidney appeared to be the major organ affected. 

In the 6-month study, Fischer 344 rats (10/sex /group) were administered 0, 0.312, 0.625, 1.25, 2.5 or 5

mg/kg-day of HgCl2 (0.23, 0.46, 0.92, 1.9 and 3.7 mg/kg-day) 5 days/week by gavage. Survival was not

affected, although body-weight gains were decreased in males at high dose and in females at or above the

0.46 mg/kg-day dose. Absolute and relative kidney weights were significantly increased in both sexes

with exposure to at least 0.46 mg/kg-day. In males, the incidence of nephropathy was 80% in the

controls and 100% for all treated groups; however, severity was minimal in the controls and two

low-dose groups and minimal to mild in the 0.92 mg/kg-day group and higher. In females, there was a

significant increased incidence of nephropathy only in the high-dose group (4/10 with minimal severity). 

Nephropathy was characterized by foci of tubular regeneration, thickened tubular basement membrane

and scattered dilated tubules containing hyaline casts. No treatment-related effects were 

observed in the other organs; however, histopathology on the other organs was performed only on control

and high-dose rats.

B6C3F1 mice (10/sex/group) were administered 0, 1.25, 2.5, 5, 10 or 20 mg/kg-day HgCl2 (0,

0.92, 1.9,3.7, 7.4 or 14.8 mg/kg-day) 15 days/week by gavage for 6 months (NTP 1993). A decrease in

body-weight gain was reported in only the males at the highest dose tested. Significant increases

occurred in absolute kidney weights of male mice at 3.7 mg/kg-day or greater and relative kidney weights

of male mice at 7.4 and 14.8 mg/kg-day doses. The kidney weight changes corresponded to an increased

incidence of cytoplasmic vacuolation of renal tubule epithelium in males exposed to at least 3.7

mg/kg-day. The exposed female mice did not exhibit any histopathologic changes in the kidneys. 

In the 2-year NTP study, Fischer 344 rats (60/sex/group) were administered 0, 2.5 and 5

mg/kg-day HgCl2 (1.9 and 3.7 mg/kg-day) 5 days week by gavage (NTP, 1993). After 2 years, survival

was reduced in only the treated male rat groups compared with the control. Mean body weights were

decreased in both male and female treated groups. After 2 years, an increased incidence of nephropathy

of moderate-to-marked severity and increased incidence of tubule 

hyperplasia was observed in the kidneys of exposed males compared with the controls. The control

males exhibited nephropathy, primarily of mild-to-moderate severity. Hyperparathyroidism,

mineralization of various tissues and fibrous osteodystrophy were observed and considered secondary to

the renal impairment. No significant differences were found in renal effects between exposed and

control females. Other nonneoplastic effects included an increased incidence of forestomach hyperplasia

in the exposed males and high-dose females. 

NTP (1993) also administered to B6C3F1 mice (60/sex/group) daily oral gavage doses of 0, 5 or

10 mg/kg-day HgCl2 (0, 3.7 and 7.4 mg/kg-day) 5 days/week by gavage for 2 years. Survival and body

weights of mice were slightly lower in HgCl2-treated mice compared with controls. Absolute kidney

weights were significantly increased in the treated males, while relative kidney weights were

significantly increased in high-dose males and both low- and high-dose females. Histopathology

revealed an increase in the incidence and severity of nephropathy in exposed males and an increase in the

incidence of nephropathy in exposed females. Nephropathy was defined as foci of proximal convoluted

tubules with thickened basement membrane and basophilic cells with scant cytoplasm. Some affected
B-25

convoluted tubules contained syaline casts. Also, an increase in nasal cavity inflammation (primarily

infiltration of granulocytes in nasal mucosa) was observed in the exposed animals. 

Gale and Ferm (1971) studied the teratogenic effects of mercuric acetate on Syrian golden

hamsters. Single doses of 2, 3 or 4 mg/kg were injected by the i.v. route on day 8 of gestation. Growth

retardation, increased resorption rates and edema of the fetuses were found at all three dose levels, while

an increase in the number of abnormalities was detected at the two higher doses. In a more recent study,

Gale (1981) compared the embryotoxic effects of a single s.c. dose of 15 mg/kg mercuric acetate on the

eighth day of gestation in five inbred strains and one noninbred strain of Syrian hamsters. While strain

differences were apparent, a variety of abnormalities were reported in all the strains. Gale (1974) also

compared the relative effectiveness of different exposure routes in Syrian hamsters. The following

sequence of decreasing efficacy was noted for mercuric acetate; i.p. > i.v. > s.c. > oral. The lowest doses

used, 2 mg/kg for i.p. and 4 mg/kg for the other three routes, were all effective in causing increased

resorption and percent abnormalities. 

In male mice administered a single i.p. dose of 1 mg/kg HgCl2, fertility decreased between days

28 and 49 post treatment with no obvious histological effects noted in the sperm (Lee and Dixon, 1975). 

The period of decreased fertility indicated that spermatogonia and premeiotic spermatocytes were

affected. The effects were less severe than following a similar dose of methyl mercury. A single i.p.

dose of 2 mg/kg HgCl2 in female mice resulted in a significant decrease in the total number of implants

and number of living embryos and a significant increase in the percentage of dead implants (Suter, 1975). 

These effects suggest that mercury may be a weak inducer of dominant lethal mutations.

 I.A.5 CONFIDENCE IN THE ORAL RfD

Study -- N/A

Data Base -- High

RfD -- High

No one study was found adequate for deriving an oral RfD; however, based on the weight of

evidence from the studies using Brown Norway rats and the entirety of the mercuric mercury data base,

an oral RfD of high confidence results. 

 I.A.6 EPA DOCUMENTATION AND REVIEW OF THE ORAL RfD

Source Document -- U.S. EPA, 1988

This IRIS summary is included in The Mercury Study Report to Congress, which was reviewed

by OHEA and EPA's Mercury Work Group in November 1994. An interagency review by scientists from

other federal agencies took place in January 1995. The report was also reviewed by a panel of

non-federal external scientists in January 1995 who met in a public meeting on January 25-26. All

reviewers comments have been carefully evaluated and considered in the revision and finalization of this

IRIS summary. A record of these comments is 

summarized in the IRIS documentation files.
B-26

Other Docmentation -- U.S. EPA, 1987

Agency Work Group Review -- 08/05/85, 02/05/86, 08/19/86, 11/16/88 

Verification Date -- 11/16/88

 I.A.7 EPA CONTACTS (ORAL RfD)

W. Bruce Peirano / NCEA -- (513)569-7540

Krishan Khanna / OST -- (202)260-7588

REFERENCES

Andres, P. 1984. IgA-IgG disease in the intestine of Brown Norway rats ingesting mercuric chloride. 

Clin. Immunol. Immunopathol. 30: 488-494. 

Bernaudin, J.F., E. Druet, P. Druet and R. Masse. 1981. Inhalation or ingestion of organic or inorganic

mercurials produces auto-immune disease in rats. Clin. Immunol. Immunopathol. 20: 129-135. 

Druet, P., E. Druet, F. Potdevin and C. Sapin. 1978. Immune type glomerulonephritis induced by HgCl2

in the Brown Norway rat. Ann. Immunol. 129C: 777-792. 

Fitzhugh, O.G., A.A. Nelson, E.P. Laug and F.M. Kunze. 1950. Chronic oral toxicants of

mercuric-phenyl and mercuric salts. Arch. Ind. Hyg. Occup. Med. 2: 433-442. 

Gale, T.F. 1974. Embryopathic effects of different routes of administration of mercuric acetate in the

hamster. Environ. Res. 8: 207-213. 

Gale, T.F. 1981. The embryotoxic response produced by inorganic mercury in different strains of

hamsters. Environ. Res. 24: 152-161. 

Gale, T. and V. Ferm. 1971. Embryopathic effects of mercuric salts. Life Sci. 10(2): 1341-1347. 

Kazantzis, G., K.F.R. Schiller, A.W. Asscher and R.G. Drew. 1962. Albuminuria and the nephrotic

syndrome following exposure to mercury and its compounds. Q. J. Med. 31(124): 403-419. 

Lee, I.D. and R.L. Dixon. 1975. Effects of mercury on spermatogenesis studied by velocity

sedimentation, cell separation and serial mating. J. Pharmacol. Exp. Ther. 194(1): 171-181. 

NTP (National Toxicology Program). 1993. Toxicology and carcinogenesis studies of mercuric chloride

(CAS No. 7487-94-7) in F344 rats and B3C3F1 mice (gavage studies). NTP Technical Report Series No.

408. National Toxicology Program, U.S. Department of Health and Human Services, Public Health

Service, National Institutes of Health, Research Triangle Park, NC. 
B-27

Suter, K.E. 1975. Studies on the dominant lethal and fertility effects of he heavy metal compounds

methyl mercuric hydroxide, mercuric chloride, and cadmium chloride in male and female mice. Mutat.

Res. 30: 365-374. 

U.S. EPA. 1987. Peer Review Workshop on Mercury Issues. Environmental Criteria and Assessment

Office, Cincinnati, OH. Summary report. October 26-27. 

U.S. EPA. 1988. Drinking Water Criteria Document for Inorganic Mercury. Prepared by the Office of

Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH

for the Office of Drinking Water, Washington, DC. EPA/600/X-84/178. NTIS PB89-192207. 
B-28

 II. CARCINOGENICITY ASSESSMENT FOR LIFETIME EXPOSURE

Substance Name -- Mercuric Chloride

CASRN -- 7487-94-7

Preparation Date -- 5/24/94

 II.A EVIDENCE FOR CLASSIFICATION AS TO HUMAN CARCINOGENICITY

 II.A.1 WEIGHT-OF-EVIDENCE CLASSIFICATION

Classification -- C; possible human carcinogen

Basis -- Based on the absence of data in humans and limited evidence of carcinogenicity in rats and mice. 

Focal papillary hyperplasia and squamous cell papillomas in the forestomach as well as thyroid follicular

cell adenomas and carcinomas were observed in male rats gavaged with mercuric chloride for 2 years. 

The relevance of the forestomach papillomas to assessment of cancer in humans is questionable because

no evidence indicated that the papillomas progressed to malignancy. The relevance of the increase in

thyroid tumors has also been questioned because these tumors are generally considered to be secondary

to hyperplasia; this effect was not observed in the high-dose males. It should also be noted that the

authors considered the doses used in the study to exceed the MTD for male rats. In the same study,

evidence for increases in squamous cell papillomas in the forestomach of female rats was equivocal. In a

second study, equivocal evidence for renal adenomas and adenocarcinomas was observed in male mice;

there was a significant positive trend. This tumor type is rare in mice, and the increase in incidence was

statistically significant when compared with historic controls. Two other nonpositive lifetime rodent

studies were considered inadequate. Mercuric chloride showed mixed results in a number of

genotoxicity assays.

 II.A.2 HUMAN CARCINOGENICITY DATA

None. No data are available on the carcinogenic effects of mercuric chloride in humans. 

 II.A.3 ANIMAL CARCINOGENICITY DATA

Limited. The results from a dietary study in rats and mice show equivocal evidence for

carcinogenic activity in male mice and female rats and some evidence for carcinogenic activity in male

rats. Two other dietary studies did not show any evidence for carcinogenicity, but these studies are

limited by inadequacies in the data and experimental design, including the small number of animals/dose

and/or a lack of complete histopathological examinations. 

Mercuric chloride (purity >99%) was administered by gavage in water at doses of 0, 2.5 or 5

(mg/kg)/day (0, 1.9 and 3.7 (mg/kg)/day) to 60 F344 rats/sex/group, 5 days/week for 104 weeks (NTP,
B-29

1993). An interim sacrifice (10/sex/dose) was conducted after 15 months of exposure. Complete

histopathological examinations were performed on all animals found dead, killed in extremis, or killed by

design. Survival after 24 months was lower in low- and high-dose males at a statistically significant rate;

survival was 43, 17 and 8% in control, low-, and high-dose males, respectively, and 58, 47 and 50% in

control, low-, and high-dose females, respectively. During the second year of the study, body weight

gains of low- and high-dose males were 91 and 85% of controls, respectively, and body weight gains of

low- and high-dose females were 90 and 86% of controls, respectively. At study termination,

nephropathy was evident in almost all male and female rats including controls, but the severity was much

greater in treated males. The incidence of "marked" nephropathy was 6/50, 29/50 and 29/50 in control,

low- and high-dose males, respectively. Squamous cell papillomas of the forestomach showed a

statistically significant positive trend with dose by life table adjusted analysis; the incidences were 0/50,

3/50 and 12/50 in control, low- and high-dose males, respectively. For females, the incidence was 0/50,

0/49 and 2/50 in control, low- and high-dose groups, respectively. These neoplasms are rare in male rats

and occurred in only 1/264 historical controls. The incidence of papillary hyperplasia of the stratified

squamous epithelium lining of the forestomach was elevated at a statistically significant rate in all dosed

males (3/49, 16/50 and 35/50 in control, low- and high-dose males, respectively) and in high-dose

females (5/50, 5/49 and 20/50 in control, low-and high-dose females, respectively). The incidence of

thyroid follicular cell carcinomas, adjusted for survival, showed a significant positive trend in males; the

incidence was 1/50, 2/50 and 6/50 in control, low- and high-dose groups, respectively. The combined

incidence of thyroid follicular cell neoplasms (adenoma and/or carcinoma) was not significantly

increased (2/50, 6/50 and 6/50 in control, low- and high-dose males, respectively). In female rats a

significant decrease in the incidence of mammary gland fibroadenomas was observed (15/50, 5/48 and

2/50 in control, low- and high-dose females, respectively). The high mortality in both groups of treated

males indicates that the MTD was exceeded in these groups and limits the value of the study for

assessment of carcinogenic risk. NTP (1993) considered the forestomach tumors to be of limited

relevance to humans because the tumors did not appear to progress to malignancy. NTP (1993) also

questioned the relevance of the thyroid carcinomas because these neoplasms are usually seen in

conjunction with increased incidences of hyperplasia and adenomas. In this study, however, no 

increases in hyperplasia or adenomas were observed. Hyperplasia incidence was 2/50, 4/50 and 2/50 in

control, low- and high-dose males, respectively; adenoma incidence was 1/50, 4/50 and 0/50 in control,

low- and high-dose males, respectively. 

In the same study, mercuric chloride was administered by gavage in water at doses of 0, 5 or 10

(mg/kg)/day (0, 3.7 and 7.4 (mg/kg)/day) to 60 B6C3F1 mice/sex/group 5 days/week for 104 weeks

(NTP, 1993). An interim sacrifice (10/sex/dose) was conducted after 15 months of exposure. Terminal

survival and body weight gain were not affected in either sex by the administration of mercuric chloride. 

It should be noted that survival of high-dose females was lower than controls; female survival rates were

82, 70 and 62% in control, low- and high-dose females, respectively. Female mice exhibited a significant

increase in the incidence of nephropathy (21/49, 43/50 and 42/50 in control, low- and high-dose females,

respectively). Nephropathy was observed in 80-90% of the males in all groups. The severity of

nephropathy increased with increasing dose. The incidence of renal tubular hyperplasia was 1/50, 0/50

and 2/49 in control, low- and high-dose males. The combined incidence of renal tubular adenomas and

adenocarcinomas was 0/50, 0/50 and 3/49 in control, low- and high-dose males, respectively. Although

no tumors were seen in the low-dose males, a statistically significant positive trend for increased

incidence with increased dose was observed. These observations were considered important because

renal tubular hyperplasia and tumors in mice are rare. The 2-year historical incidence of renal tubular

adenomas or adenocarcinomas in males dosed by gavage with water was 0/205, and only 4 of the nearly

400 completed NTP studies have shown increased renal tubular neoplasms in mice. Data from this study
B-30

were not statistically compared with historical control data by NTP. EPA's analysis of the reported data

with Fisher's Exact test showed that the incidence of renal tubular adenomas or adenocarcinomas in the

high-dose males was significantly elevated when compared with historical controls (Rice and Knauf,

1994).

A 2-year feeding study in rats (20 or 24/sex/group; strain not specified) was conducted in which

mercuric acetate was administered in the diet at doses of 0, 0.5, 2.5, 10, 40 and 160 ppm (0, 0.02, 0.1,

0.4, 1.7 and 6.9 (mg Hg/kg)/day (Fitzhugh et al., 1950). Survival was not adversely affected in the study. 

Increases in kidney weight and renal tubular lesions were observed at the two highest doses. No

statement was made in the study regarding carcinogenicity. This study was not intended to be a

carcinogenicity assay, and the number of animals/dose was rather small. Histopathological analyses

were conducted on only 50% of the animals (complete histopathology conducted on only 31% of the

animals examined), and no quantitation of results or statistical analyses were performed.

No increase in tumor incidence was observed in a carcinogenicity study using white Swiss mice

(Schroeder and Mitchener, 1975). Groups of mice (54/sex/group) were exposed until death to mercuric

chloride in drinking water at 5 ppm Hg (0.95 (mg/kg)/day). No effects on survival or body weights were

observed. After dying, mice were weighed and dissected. The animals were examined for gross tumors,

and some sections were made of the heart, lung, liver, kidney and spleen for microscopic examination. 

No toxic effects of mercuric chloride were reported in the study. No statistically significant differences

were observed in tumor incidences for treated animals and controls. This study is of limited use for

evaluation of carcinogenicity because complete histological examinations were not performed, only a

single dose was tested, and the MTD was not achieved.

 II.A.4 SUPPORTING DATA FOR CARCINOGENICITY

The increasing trend for renal tubular cell tumors in mice observed in the NTP (1993) study

receives some support from similar findings in mice after chronic dietary exposure to methylmercury

(Hirano et al., 1986; Mitsumori et al., 1981, 1990). In these studies, dietary exposure to methylmercuric

chloride resulted in increases in renal tubular tumors at doses wherein substantial nephrotoxicity was

observed (see methylmercury file on IRIS).

As summarized in NTP (1993) and U.S. EPA (1985), mercuric chloride has produced some

positive results for clastogenicity in a variety of in vitro and in vivo genotoxicity assays; mixed results

regarding its mutagenic activity have been reported. Mercuric chloride was negative in gene mutation

tests with Salmonella typhimurium (NTP, 1993; Wong, 1988) but produced DNA damage as measured in

the Bacillus subtilis rec assay (Kanematsu et al., 1980). A 

weakly positive response for gene mutations was observed in mouse lymphoma (L5178Y) cells in the

presence of microsomal activation (Oberly et al., 1982). DNA damage has also been observed in assays

using rat and mouse embryo fibroblasts (Zasukhina et al., 1983), CHO cells and human KB cells

(Cantoni and Costa, 1983; Cantoni et al., 1982, 1984a,b; Christie et al., 1984, 1986; 

NTP, 1993; Williams et al., 1987). Mercuric chloride also produced chromosome aberrations and SCEs

in CHO cells (Howard et al., 1991) and chromosome aberrations in human lymphocytes (Morimoto et al.,

1982). Sex-linked recessive lethal mutations were not observed in male Drosophila melanogaster (NTP,

1993).
B-31

Although mice given intraperitoneal doses of mercuric chloride have shown no increase in

chromosomal aberrations in bone marrow cells (Poma et al., 1981) and no increase in aneuploidy in

spermatogonia (Jagiello and Lin, 1973), mercuric chloride administered to mice by gavage induced a

dose-related increase in chromosome aberrations and aberrant cells in the bone marrow (Ghosh et al.,

1991). Similarly, an increased incidence of chromosomal aberrations (primarily deletion and numeric

aberrations) was observed in livers of fetal mice exposed to mercury in utero as the result of maternal

inhalation of aerosols of mercuric chloride (Selypes et al., 1984). Positive dominant lethal results

(increased resorptions and post-implantation deaths in 

untreated females) have been obtained in studies in which male rats were administered mercuric chloride

orally (Zasukhina et al., 1983). A slight increase in post-implantation deaths and a decrease in living

embryos were also reported in treated female mice mated to untreated males (Suter, 1975); however, it

was not clear whether these effects were the result of germ cell 

mutations or were secondary to maternal toxicity.

The effects of mercuric chloride on genetic material has been suggested to be due to the ability of

mercury to inhibit the formation of the mitotic spindle, an event known as c-mitosis (U.S. EPA, 1985).

 II.B QUANTITATIVE ESTIMATE OF CARCINOGENIC RISK FROM ORAL EXPOSURE

 None. The incidences of squamous cell papillomas of the forestomach and thyroid follicular

cell carcinomas were evaluated. No slope factor was derived using the forestomach tumors because these

tumors are probably the result of doses of mercuric chloride above-MTD resulting in irritation of the

forestomach and subsequent cell death and epithelial proliferation. The carcinogenic mechanism for

mercuric chloride at the high doses observed may be specific to effects of irritation of the forestomach.

Regarding the thyroid carcinomas, a variety of drugs, chemicals and physiological perturbations

result in the development of thyroid follicular tumors in rodents. For a number of chemicals, the

mechanism of tumor development appears to be a secondary effect of long-standing hypersecretion of

thyroid-stimulating hormone by the pituitary (Capen and Martin, 1989; McClain, 1989). In the absence

of such long-term stimulatory effects, induction of thyroid follicular cell cancer by such chemicals

usually does not occur (Hill, 1989). The mechanism whereby thyroid tumors developed in the NTP

(1993) assay is very unclear given that hyperplasia was not observed. The study reviewers concluded

that it was difficult to associate the increase in thyroid tumors with mercuric chloride administration. 

Thus, it would be of questionable value to use the thyroid tumors in rats as the basis for a quantitative

cancer risk estimate for humans. 

All tumors in rats were observed at doses equalling or exceeding the MTD. Kidney tumors in

mice were observed in only the high-dose males. The increased incidence was not statistically significant

in comparison to the concurrent controls, but was significant when compared with historical controls. A

linear low-dose extrapolation based on the male mouse kidney tumor data (three tumors in the high-dose

group only) is not appropriate. 

 II.C QUANTITATIVE ESTIMATE OF CARCINOGENIC RISK FROM INHALATION

EXPOSURE

None.
B-32

 II.D EPA DOCUMENTATION, REVIEW AND CONTACTS (CARCINOGENICITY

ASSESSMENT)

 II.D.1 EPA DOCUMENTATION

Source Document -- U.S. EPA, 1995

This IRIS summary is included in The Mercury Study Report to Congress which was reviewed

by OHEA and EPA's Mercury Work Group in November 1994. An interagency review by scientists from

other federal agencies took place in January 1995. The report was also reviewed by a panel of

non-federal external scientists in January 1995 who met in a public meeting on January 25-26. All

reviewers comments have been carefully evaluated and considered in the revision and finalization of this

IRIS summary. A record of these comments is summarized in the IRIS documentation files.

 II.D.2 REVIEW (CARCINOGENICITY ASSESSMENT)

Agency Work Group Review -- 03/03/94 

Verification Date -- 03/03/94

 II.D.3 U.S. EPA CONTACTS (CARCINOGENICITY ASSESSMENT)

Rita Schoeny / NCEA -- (513)569-7544

REFERENCES

Cantoni, O. and M. Costa. 1983. Correlations of DNA strand breaks and their repair with cell survival

following acute exposure to mercury (II) and X-rays. Mol. Pharmacol. 24(1): 84-89. 

Cantoni, O., R.M. Evans and M. Costa. 1982. Similarity in the acute cytotoxic response of mammalian

cells to mercury (II) and X-rays: DNA damage and glutathione depletion. Biochem. Biophys. Res.

Commun. 108(2): 614-619.

Cantoni, O., N.T. Christie, A. Swann, D.B. Drath and M. Costa. 1984a. Mechanism of HgCl2

cytotoxicity in cultured mammalian cells. Mol. Pharmacol. 26: 360-368. 

Cantoni, O., N.T. Christie, S.H. Robinson and M. Costa. 1984b. Characterization of DNA lesions

produced by HgCl2 in cell culture systems. Chem. Biol. Interact. 49: 209-224. 

Capen, C.C. and S.L. Martin. 1989. The effects of xenobiotics on the structure and function of thyroid

follicular and C-cells. Toxicol. Pathol. 17(2): 266-293. 
B-33

Christie, N.T., O. Cantoni, R.M. Evans, R.E. Meyn and M. Costa. 1984. Use of mammalian DNA

repair-deficient mutants to assess the effects of toxic metal compounds on DNA. Biochem. Pharmacol. 

33(10): 1661-1670. 

Christie, N.T., O. Cantoni, M. Sugiyama, F. Cattabeni and M. Costa. 1986. Differences in the effects of

Hg(II) on DNA repair induced in Chinese hamster ovary cells by ultraviolet or X-rays. Mol. Pharmacol. 

29: 173-178.

Fitzhugh, O.G., A.A. Nelson, E.P. Lauge and F.M. Kunze. 1950. Chronic oral toxicities of

mercuric-phenyl and mercuric salts. Arch. Ind. Hyg. Occup. Med. 2: 433-442. 

Ghosh, A.K., S. Sen, A. Sharma and G. Talukder. 1991. Effect of chlorophyllin on mercuric

chloride-induced clastogenicity in mice. Food. Chem. Toxicol. 29(11): 777-779. 

Hill, R.N., L.S. Erdreich, O.V. Paynter, P.A. Roberts, S.L. Rosenthal and C.F. Wilkinson. 1989. 

Review. Thyroid follicular cell carcinogenesis. Fund. Appl. Toxicol. 12: 629-697. 

Hirano, M., K. Mitsumori, K. Maita and Y. Shiraso. 1986. Further carcinogenicity study on

methylmercury chloride in ICR mice. Jap. J. Vet. Sci. 48(1): 127-135. 

Howard, W., B. Leonard, W. Moody and T.S. Kochhar. 1991. Induction of chromosome changes by

metal compounds in cultured CHO cells. Toxicol. Lett. 56(1-2): 179-186. 

Jagiello, G. and J.S. Lin. 1973. An assessment of the effects of mercury on the meiosis of mouse ova. 

Mutat. Res. 17: 93-99.

Kanematsu, N., M. Hara and T. Kada. 1980. Rec assay and mutagencity studies on metal compounds. 

Mutat. Res. 77: 109-116.

McClain, R.M. 1989. The significance of hepatic microsomal enzyme induction and altered thyroid

function in rats: Implications for thyroid gland neoplasia. Toxicol. Pathol. 17(2): 294-306. 

Mitsumori, K., K. Maita, T. Saito, S. Tsuda and Y. Shirasu. 1981. Carcinogenicity of methylmercury

chloride in ICR mice: Preliminary note on renal carcinogenesis. Cancer Lett. 12: 305-310.

Mitsumori, K., M. Hirano, H. Ueda, K. Maita and Y. Shirasu. 1990. Chronic toxicity and

carcinogenicity of methylmercury chloride in B6C3F1 mice. Fund. Appl. Toxicol. 14: 179-190. 

Morimoto, K., S. Iijima and A. Koizumi. 1982. Selenite prevents the induction of sister-chromatid

exchanges by methyl mercury and mercuric chloride in human whole-blood cultures. Mutat. Res. 102:

183-192. 

NTP (National Toxicology Program). 1993. NTP technical report on the toxicology and carcinogenesis

studies of mercuric chloride (CAS No. 7487-94-7) in F344 rats and B6C3F1 mice (gavage studies). NTP

TR 408. National Toxicology Program, U.S. Department of Health and Human Services, Public Health

Service, National Institutes of Health, Research Triangle Park, NC.
B-34

Oberly, T.J., C.E. Piper and D.S. McDonald. 1982. Mutagenicity of metal salts in the L5178Y mouse

lymphoma assay. J. Toxicol. Environ. Health. 9: 367-376. 

Poma, K., M. Kirsch-Volders and C. Susanne. 1981. Mutagenicity study of mice given mercuric

chloride. J. Appl. Toxicol. 1(6): 314-316.

Rice, G. and L. Knauf. 1994. Further Statistical Evaluation of the NTP Mercuric Chloride Mouse

Bioassay. Memorandum to the U.S. EPA CRAVE File for Mercuric Chloride, March 1.

Schroeder, H. and M. Mitchener. 1975. Life-time effects of mercury, methyl mercury, and nine other

trace metals in mice. J. Nutr. 105: 452-458.

Selypes, A., L. Nagymajtenyi and G. Berencsi. 1984. Study of the mutagenic and teratogenic effect of

aerogenic mercury exposition in mouse. Collect. Med. Leg. Toxicol. Med. 125: 65-69. 

Suter, K.E. 1975. Studies on the dominant-lethal and fertility effects of the heavy metal compounds

methylmercuric hydroxide, mercuric chloride and cadmium chloride in male and female mice. Mutat.

Res. 30: 365-374.

U.S. EPA. 1980. Ambient Water Quality Criteria Document for Mercury. Prepared by the Office of

Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH

for the Office of Water Regulation and Standards, Washington, DC. EPA/440/5-80/058. NTIS PB

81-117699. 

U.S. EPA. 1984a. Mercury Health Effects Update: Health Issue Assessment. Final Report. Prepared

by the Office of Health and Environmental Assessment, Environmental Criteria and Assessment Office,

Cincinnati, OH for the Office of Air Quality Planning and Standards, Research Triangle Park, NC. 

EPA/600/8-84/019F. NTIS PB81-85-123925. 

U.S. EPA. 1984b. Health Effects Assessment for Mercury. Prepared by the Office of Health and

Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH for the

Office of Emergency and Remedial Response, Washington, DC. EPA/540/1086/042. NTIS

PB86-134533/AS. 

U.S. EPA. 1985. Drinking Water Criteria Document for Mercury. Prepared by the Office of Health and

Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH for the

Office of Drinking Water, Washington, DC. EPA/600/X-84/178. NTIS PB86-117827. 

U.S. EPA. 1988. Drinking Water Criteria Document for Inorganic Mercury. Prepared by the Office of

Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH

for the Office of Drinking Water, Washington, DC. EPA/600/X-84/178. NTIS PB89-192207.

U.S. EPA. 1993. Summary Review of Health Effects Associated with Mercuric Chloride: Health Issue

Assessment (Draft). Prepared by the Office of Health and Environmental Assessment, Environmental

Criteria and Assessment Office, Cincinnati, OH for the Office of Air Quality Planning and Standards,

Research Triangle Park, NC. EPA/600/R-92/199. 

U.S. EPA. 1995. Mercury Study Report to Congress. Office of Research and Development,

Washington, DC. External Review Draft. EPA/600/P-94/002Ab.
B-35

Williams, M.V., T. Winters and K.S. Waddel. 1987. In vivo effects of mercury (II) on deoxyuridine

triphosphate nucleotidohydrolase, DNA polymerase (alpha, beta) and uracil-DNA glycosylase activities

in cultured human cells: Relationship to DNA damage, DNA repair, and cytotoxicity. Mol. Pharmacol. 

31: 200-207.

Wong, P.K. 1988. Mutagenicity of heavy metals. Bull. Environ. Contam. Toxicol. 40(4): 597-603.

Zasukhina, G.D., I.M. Vasilyeva, N.I. Sdirkova, G.N. Krasovsky, L.Y. Vasyukovich, U.I. Kenesariev and

P.G. Butenko. 1983. Mutagenic effect of thallium and mercury salts on rodent cells with different repair

activities. Mutat. Res. 124: 163-173. 
B-36

 I.A. REFERENCE DOSE FOR CHRONIC ORAL EXPOSURE (RfD)

Chemical -- Methylmercury (MeHg)

CASRN -- 22967-92-6

Preparation Date -- 2/10/95

 I.A.1 ORAL RfD SUMMARY

Critical Effect Experimental Doses* UF MF RfD

Developmental Benchmark Dose: 11 ppm 10 1 1E-4

neurologic in hair; equivalent to mg/kg-day

abnormalities maternal blood levels

in human infants 44 ug/L and body

burdens of 69 ug or

Human epidemiologic daily intake of 1.1

studies ug/kg-day

Marsh et al., 1987; Seafood Safety, 1991

* Conversion Factors and Assumptions -- Maternal daily dietary intake levels were used as the dose

surrogate for the observed developmental effects in the infants. The daily dietary intake levels were

calculated from hair concentrations measured in the mothers. This conversion is explained in the text

below. A benchmark dose approach was used rather than a NOAEL/LOAEL approach to analyze the

neurological effects in infants as the response variable. This analysis is also explained in the text below.

 I.A.2 PRINCIPAL STUDIES (ORAL RfD)

Marsh, D.O., T.W. Clarkson, C. Cox, L. Amin-Zaki and S. Al-Trkiriti. 1987. Fetal methylmercury

poisoning: Relationship between concentration in a single strand of maternal hair and child effects. 

Arch. Neurol. 44: 1017-1022. 

Seafood Safety. 1991. Committee on Evaluation of the Safety of Fishery Products, Chapter on

Methylmercury: FDA Risk Assessment and Current Regulations, National Academy Press, Washington,

DC. p. 196-221. 

In 1971-1972 many citizens in rural Iraq were exposed to MeHg-treated seed grain that was

mistakenly used in home-baked bread. Latent toxicity was observed in many adults and children who

had consumed bread over a 2- to 3-month period. Infants born to mothers who ate contaminated bread

during gestation were the most sensitive group. Often infants exhibited neurologic abnormalities while

their mothers showed no signs of toxicity. Some information indicates that male infants are more

sensitive than females. Among the signs noted in the infants exposed during fetal development were

cerebral palsy, altered muscle tone and deep tendon reflexes as well as delayed developmental

milestones, i.e., walking by 18 months and talking by 24 months. The neurologic signs noted in adults
B-37

included paresthesia, ataxia, reduced visual fields and hearing impairment. Some mothers experienced

paresthesia and other sensory disturbances but these symptoms were not necessarily correlated with

neurologic effects in their children. Unique analytic features of mercury (Hg), that is, analysis of

segments of hair correlated to specific time periods in the past permitted approximation of maternal

blood levels that the fetuses were exposed to in utero. The data collected by Marsh et al. (1987)

summarizes clinical neurologic signs of 81 mother and child pairs. From x-ray fluorescent spectrometric

analysis of selected regions of maternal scalp hair, concentrations ranging from 1 to 674 ppm were

determined and correlated with clinical signs observed in the affected members of the mother-child pairs. 

Among the exposed population were affected and unaffected individuals throughout the dose-exposure

range.

While the purpose of the Seafood Safety publication was to critique the quantitative risk

assessment that FDA had performed for MeHg, this material is included in the EPA risk assessment

because the Tables of Incidence of various clinical effects in children that were provided in the FDA

assessment readily lend themselves to a benchmark dose approach. Specifically the continuous data for

the Iraqi population that was reported in Marsh et al. (1987) are placed in five dose groups and incidence

rates are provided for delayed onset of walking, delayed onset of talking, mental symptoms, seizures,

neurological scores above 3 and neurological scores above 4 for affected children. Neurologic scores

were determined by clinical evaluation for cranial nerve signs, speech, involuntary movement, limb tone

strength, deep tendon reflexes, plantar responses, coordination, dexterity, primitive reflexes, sensation,

posture, and ability to sit, stand and run. This paper provided groupings of the 81 mother-infant pairs for

various effects, and the authors present the data in Tables 6-11 through 6-16B.

EQUATION USED FOR CALCULATION OF DAILY DOSE: From the concentration of Hg present in

maternal hair, a corresponding blood concentration value is determined. A hair concentration of 11 ppm

converts to a blood concentration of 44 ug/L; the following equation can then be used to determine the

daily dose that corresponds to that blood concentration of Hg. Use of this equation is based on the

assumption that steady-state conditions exist and that firstorder kinetics for Hg are being followed.

d = (C x b x V)/(A x f)

d (ug/day) = 44 ug/L multiplied by 0.014 multiplied by 5 liters divided by 0.95 then divided by

0.05 yields 65 ug/day

where:

d = daily dietary intake (expressed as ug of MeHg)

C = concentration in blood (expressed as ug/L)

b = elimination constant (expressed as days-1)

V = volume of blood in the body (expressed as liters)

A = absorption factor (unitless)

f = fraction of daily intake taken up by blood (unitless)
B-38

The following sections provide the data and rationale supporting the choice of parameter values

used in the conversion equation. It should be noted that even if the upper or lower ranges of the

parameter values were used, the conversion factor precision remains the same due to rounding error. The

Agency realizes that new pharmacokinetic data may become available that warrant a change to some of

these parameters. 

HAIR TO BLOOD CONCENTRATION RATIO: The hair:blood concentration ratio for total Hg is

frequently cited as 250. The following description provides a justification of why we have chosen to use

the ratio of 250:1. Ratios reported in the literature range from 140 to 370, a difference of more than a

factor of 2.5. Differences in the location of hair sampled (head vs chest and distance from scalp) may

contribute to the differences observed. As much as a 3-fold seasonal variation in Hg levels was observed

in average hair levels for a group of individuals with moderate to high fish consumption rates, with

yearly highs occurring in the fall and early winter (Phelps et al., 1980; Suzuki et al., 1993). The high

slope reported by Tsubaki and Irukayama (1977) may have reflected the fact that Hg levels were

declining at the time of sampling so that the hair levels reflect earlier, higher blood levels. Phelps et al.

(1980) obtained multiple blood samples and sequentially analyzed lengths of hair from individuals. Both

hair and blood samples were taken for 339 individuals in Northwestern Ontario. After reviewing the

various reports for converting hair concentrations to blood concentrations, the Phelps paper was selected

because of the large sample size and the attention to sampling and analysis. The ratio Phelps observed

between the total Hg concentration in hair taken close to the scalp and simultaneous blood sampling for

this group was 296. To estimate the actual ratio, the authors assumed that blood and hair samples were

taken following complete cessation of MeHg intake. They also assumed a half-life of MeHg in blood of

52 days and a lag of 4 weeks for appearance of the relevant level in hair at the scalp. Phelps also

determined that 94% of the Hg in hair was MeHg. Based on these assumptions, they calculated that if

the actual hair:blood ratio was 200, they would have observed a ratio of 290. Based on these and other

considerations, Phelps states that the actual ratio is "probably higher than 200, but less than the observed

value of 296." As the authors point out, 2/3 of the study population were sampled during the falling

phase of the seasonal variation (and 1/3 or less in the rising phase). This methodology would tend to

result in a lower observed ratio; therefore, the actual average is likely to be greater than 200. 

In view of these limitations a value of 250 was considered acceptable for the purpose of

estimating average blood levels in the Iraqi population. 

CALCULATION OF DIETARY INTAKE FROM BLOOD CONCENTRATION: The first step in this

process is to determine the fraction of Hg in diet that is absorbed. Radio-labeled methyl-mercuric nitrate

(MeHgNO3) was administered in water to three healthy volunteers (Aberg et al., 1969). The uptake was

>95%. Miettinen et al. (1971) incubated fish liver homogenate with radio-labeled MeHgNO3 to produce

methylmercury proteinate. The proteinate was then fed to fish that were killed after a week and then

cooked and fed to volunteers after confirmation of MeHg in the fish. Mean uptake exceeded 94%. 

Based on these experimental results, this derivation used an absorption factor of 0.95. 

The next step involves determining the fraction of the absorbed dose that is found in the blood. 

There are three reports on the fraction of absorbed MeHg dose distributed to blood volume in humans. 

Kershaw et al. (1980) report an average fraction of 0.059 of absorbed dose in total blood volume, based

on a study of five adult male subjects who ingested MeHg-contaminated tuna. In a group of nine male

and six female volunteers who had received 203 Hg-methylmercury in fish approximately 10% of the

total body burden was present in 1 liter of blood in the first few days after exposure, dropping to

approximately 5% over the first 100 days (Miettinen et al., 1971). In another study, an average value of
B-39

1.14% for the percentage of absorbed dose in 1 kg of blood was derived from subjects who consumed a

known amount of MeHg in fish over a 3-month period (Sherlock et al., 1982). Average daily intake in

the study ranged from 43-233 ug/day and a dose-related effect on percentage of absorbed dose was

reported that ranged from 1.03-1.26% in 1 liter of blood (each of these values should be multiplied by 5

[since there are approximately 5 liters of blood in an adult human body] to yield the total amount in the

blood compartment). The value 0.05 has been used for this parameter in the past (WHO, 1990). 

ELIMINATION CONSTANT: Based on data taken from four studies, reported clearance half-times

from blood or hair ranged from 48-65 days. Two of these studies included the Iraqi population exposed

during the 1971-1972 outbreak. The value from the Cox study (Cox et al., 1989) is derived from the

study group that included the mothers of the infants upon which this risk assessment is based. The

average elimination constant of the four studies is 0.014; the average of individual values reported for 20

volunteers ingesting from 42-233 ug Hg/day in fish for 3 months (Sherlock et al., 1982) is also 0.014.

VOLUME OF BLOOD IN THE BODY AND BODY WEIGHT: Blood volume is 7% of body weight as

has been determined by various experimental methods and there is an increase of 20 to 30% (to about 8.5

to 9%) during pregnancy (Best and Taylor, 1961). Specific data for the body weight of Iraqi women

were not found. Assuming an average body weight of 60 kg. (Snyder et al., 1981) and a blood volume of

9% of body weight during pregnancy, a blood volume of 5.4 liters is derived. 

DERIVATION OF A BENCHMARK DOSE: Benchmark dose estimates were made for excess risk

above background based on a combination of all childhood neurologic end points. This method was

chosen since the Agency felt that any childhood neurologic abnormality is considered an adverse effect

and likely to have serious sequelae lasting throughout lifetime. In addition, grouping of all neurologic

endpoints provided a much better goodness of fit of the data than when any endpoint was used

individually. The endpoints that were grouped delayed the onset of walking and talking, neurologic

scores <3, mental symptoms, and seizures. Using these data sets taken from the Seafood Safety paper,

benchmark doses at the 1, 5 and 10% incidence levels were constructed using both Weibull and

polynomial models. The Weibull model places the maximum likelihood estimate with corresponding

95% confidence level at 11 ppm of MeHg in maternal hair. The Agency decided to use the lower 95%

confidence level for the 10% incidence rate. Recent research by Faustman et al. (1994) and Allen et al.

(1994a,b) suggests that the 10% level for the benchmark dose roughly correlates with a NOAEL for

quantal developmental toxicity data. The 95% lower confidence limits on doses corresponding to the 1,

5, and 10% levels were calculated using both models and the values determined using the polynomial

model always fell within 3% of the Weibull values. For final quantitative analysis the Weibull model

was chosen because of goodness of fit of the data and because this model has been used in the past by the

Agency for developmental effects. The experience of the Agency indicates that this model performs well

when modeling for developmental effects.

 I.A.3 UNCERTAINTY AND MODIFYING FACTORS (ORAL RfD)

UF -- An uncertainty factor of 3 is applied for variability in the human population, in particular the

variation in the biological half-life of MeHg and the variation that occurs in the hair:blood ratio for Hg. 

In addition, a factor of 3 is applied for lack of a two-generation reproductive study and lack of data for

the effect of exposure duration on sequelae of the developmental neurotoxicity effects and on adult

paresthesia. The total UF is 10. 
B-40

MF -- None

 I.A.4 ADDITIONAL STUDIES/COMMENTS (ORAL RfD)

McKeown-Eyssen et al. (1983) have provided a report of neurologic abnormalities in four

communities of Cree Indians in northern Quebec. A group of 247 children first exhibited clinical signs

consistent with MeHg exposure between 12 and 30 months of age. An attempt was made to account for

possible confounding factors; the interviewers determined alcohol and tobacco consumption patterns

among the mothers of affected children. Age of the mothers and multiparity was also taken into account

in analysis of the data. The average indices of exposure were the same for boys and girls at 6 ug/g; only

6% had exposure above 20 ug/g. The prevalence of multiple abnormal neurologic findings was about 4%

for children of both sexes. The most frequently observed abnormality was delayed deep tendon reflexes;

this was seen in 11.4% of the boys and 12.2% of the girls. These investigators found that when there was

a positive association between maternal Hg exposure and abnormal neurologic signs in boys, the

incidence rate was 7.2%. The incidence rate for neurologic disorders in daughters was less and was

found to be not statistically significant. Disorders of muscle tone were usually confined to the legs. 

Persistence of the Babinski reflex and incoordination due to delayed motor development were seen with

equal frequency for both sexes. The discriminant analysis conducted for the boys to distinguish the 15

cases with abnormal muscle tone or reflexes from the 82 normal controls was unable to separate

differences between these groups based on confounding variables. The prevalence of abnormality of

muscle tone or reflexes was found to increase 7 times with each increase of 10 ug/g of the prenatal

exposure index. Although this study provides supportive data for the RfD, it is not included with the

principal studies because it was confounded by alcoholism and smoking among 

mothers.

Studies performed in New Zealand investigated the mental development of children who had

prenatal exposure to MeHg (Kjellstrom et al., 1986, 1989). A group of 11,000 mothers who regularly ate

fish were initially screened by survey and of these about 1000 had consumed fish in three meals per week

during pregnancy. Working from this large population base, 31 matched pairs were established. For

proper comparison a reference child matched for ethnic group and age of mother, child's birthplace and

birth date was identified for each high Hg child. Retrospective Hg concentrations were determined from

the scalp hair of the mothers to match the period of gestation. The average hair concentration for

high-exposure mothers was 8.8 mg/kg and for the reference group it was 1.9 mg/kg.

The children of exposed mothers were tested at 4 and 6 years of age. At 4 years of age the

children were tested using the Denver Developmental Screen Test (DDST) to assess the effects of Hg. 

This is a standardized test of a child's mental development that can be administered in the child's home. 

It consists of four major function sectors: gross motor, fine motor, language, and personal-social. A

developmental delay in an individual item is scored as 

abnormal, questionable when the child has failed in their response and at least 90% of the children can

pass this item at a younger age. The results of the DDST demonstrated 2 abnormal scores and 14

questionable scores in the high Hg-exposed group and 1 abnormal and 4 questionable scores in the

control group. Analysis of the DDST results by sector showed that developmental delays were most

commonly noted in the fine motor and language sectors but the 

differences for the experimental and control groups were not significant. The investigators noted that

differences in performance of the DDST between high Hg-exposed and referent children could be due to
B-41

confounding variables and that DDST results are highly dependent upon the age of the child. 

Standardized vision tests and sensory tests were also performed to measure 

development of these components of the nervous system. The prevalence for developmental delay in

children was 52% from high Hg mothers and 17% from mothers of the reference group. In comparison to

other studied populations, the hair Hg concentration of the mothers in this study were lower than those

associated with CNS effects in children exposed in Japan and Iraq. Results of the DDST demonstrated 2

abnormal scores and 14 questionable scores in the high 

Hg-exposed group and 1 abnormal and 4 questionable scores in the control group. Analysis of the DDST

results by sector showed that developmental delays were most commonly noted in the fine motor and

language sectors but the differences for the experimental and control groups were not significant. The

data obtained from this study is too limited for detailed dose-response analysis. The differences in

performance of the DDST between high Hg-exposed and referent children could be due to confounding

variables. DDST results are highly dependent upon the age of the child. Infants of the Hg-exposed

group more frequently had low birth weights and were more likely to be born prematurely. Use of this

study is also limited by the fact that there was only a 44% participation rate.

A second stage follow-up of the original Kjellstrom study was carried out when the children

were 6 years old to confirm or refute the developmental findings observed at age 4 (Kjellstrom et al.,

1989). In this later study the high exposure children were compared with three control groups with lower

prenatal Hg exposure. The mothers of children in two of these control groups had high fish consumption

and average hair Hg concentrations during pregnancy of 3-6 mg/kg and 0-3 mg/kg, respectively. For this

study the high exposure group was matched for maternal ethnic group, age, smoking habits, residence,

and sex of the child. For this second study, 61 of 74 high-exposure children were available for study. 

Each child was tested at age 6 with an array of scholastic, psychological, and behavioral tests which

included the Test of Language Development (TOLD), the Wechsler Intelligence Scale for Children, and

the McCarthy Scale of Children's Abilities. The results of the tests were compared between groups. 

Confounding was controlled for by using linear multiple regression analysis. A principal finding was

that normal results of the psychological test variables were influenced by ethnic background and 

social class. The high prenatal MeHg exposure did decrease performance in the tests, but it contributed

only a small part of the variation in test results. The investigation found that an average hair Hg level of

13-15 mg/kg during pregnancy was consistently associated with decreased test performance. Due to the

small size of the actual study groups it was not possible to determine if even lower exposure levels might

have had a significant effect on test 

results. The Kjellstrom studies are limited for assessing MeHg toxicity because the developmental and

intelligence tests used are not the most appropriate tests for defining the effects of MeHg. Also, greater

significance was seen in differences of cultural origins of the children than the differences in maternal

hair MeHg concentrations.

The initial epidemiologic report of MeHg poisoning involved 628 human cases that occurred in

Minamata Japan between 1953 and 1960 (Tsubaki and Irukayama, 1977). The overall prevalence rate for

the Minamata region for neurologic and mental disorders was 59%. Among this group 78 deaths

occurred and hair concentrations of Hg ranged from 50-700 ug/g. Hair Hg concentrations were

determined through the use of less precise analytic methods than were available for later studies. The

specific values derived from these studies do not contribute directly to quantitative risk assessment for

MeHg. The most common clinical signs observed in adults were paresthesia, ataxia, sensory

disturbances, tremors, impairment of hearing and difficulty in walking. This particular group of

neurologic signs has become known as "Minimata disease." Examination of the brains of severely

affected patients that died revealed marked atrophy of the brain (55% normal volume and weight) with
B-42

cystic cavities and spongy foci. Microscopically, entire regions were devoid of neurons, granular cells in

the cerebellum, golgi cells and Purkinje cells. Extensive investigations of congenital Minamata disease

were undertaken and 20 cases that occurred over a 4-year period were documented. In all instances the

congenital cases showed a higher incidence of symptoms than did their mothers. Severe disturbances of

nervous function were described and the affected offspring were very late in reaching developmental

milestones. Hair concentrations of Hg in affected infants ranged from 10 to 100 ug/g. Data on hair Hg

levels for the mothers during gestation were not available.

Rice (1989) dosed five cynomolgus monkeys (Macaca fascicularis) from birth to 7 years of age

with 50 ug/kg-day and performed clinical and neurologic examinations during the dosing period and for

an additional 6 years. As an indicator of the latent effects of MeHg, objective neurologic examinations

performed at the end of the observation period revealed insensitivity to touch and loss of tactile response. 

In addition, monkeys dosed with MeHg were clumsier and slower to react when initially placed in an

exercise cage as well as in the later stages of the observation period.

Gunderson et al. (1986) administered daily doses of 50-70 ug/kg of MeHg to 11 crab-eating

macaques (Macaca fascicularis) throughout pregnancy which resulted in maternal blood levels of

1080-1330 ug/L in mothers and 1410-1840 ug/L in the offspring. When tested 35 days after birth the

infants exhibited visual recognition deficits. 

In another study, groups of 7 or 8 female crab-eating macaques (Macaca fascicularis) were dosed

with 0.50 and 90 ug/kg-day of MeHg through four menstrual cycles (Burbacher et al., 1984). They were

mated with untreated males and clinical observations were made for an additional 4 months. Two of

seven high-dose females aborted and three did not conceive during the 4-month mating period; the other

two females delivered live infants. Two of seven females of the 50 ug/kg-day dose group aborted; the

remaining females delivered live infants. All 8 females of the control group conceived and 6 delivered

live infants. These reproductive results approached but did not reach statistical significance. 

Reproductive failure within dose groups could be predicted by blood Hg levels. The dams did not show

clinical signs of MeHg poisoning during the breeding period or gestation but when females were dosed

with 90 ug/kg-day for 1 year 4/7 did show adverse neurologic signs.

Bornhausen et al. (1980) reported a decrease in operant behavior performance in 4-month-old

rats whose dams had received 0.005 and 0.05 mg/kg-day of MeHg on days 6 through 9 of gestation. A

statistically significant effect (<0.05) was observed in offspring whose dams had received 0.01 and 0.05

mg/kg during gestation. The authors postulated that more severe effects of in utero exposure would be

seen in humans since the biological half-time of Hg in the brain of humans is 5 times longer than the rat. 

In addition, much longer in utero exposure to Hg would occur in humans since gestation is much longer

in chronologic time.

In another investigation groups of Wistar rats (50/sex/dose) were administered daily doses of 2,

10, 50 and 250 ug/kg-day of MeHg for 26 months (Munro et al., 1980). Female rats that received 25

ug/kg-day had reduced body weight gains and showed only minimal clinical signs of neurotoxicity;

however, male rats that received this dose did show overt clinical signs of neurotoxicity, had decreased

hemoglobin and hematocrit values, had reduced weight gains, and showed increased mortality. 

Histopathologic examination of rats of both sexes receiving 25 ug/kg-day revealed demyelination of

dorsal nerve roots and peripheral nerves. Males showed severe kidney damage and females had minimal

renal damage. This study showed a NOAEL of 5 ug/kg-day and a LOAEL of 25 ug/kg-day. 
B-43

A 2-year feeding study of MeHg chloride was conducted in B6C3F1 mice (60 mice/sex/group) at

doses of 0, 0.4, 2 and 10 ppm (0, 0.04, 0.17, and 0.83 mg/kg-day) to determine chronic toxicity and

possible carcinogenic effects (Mitsumori et al., 1990). The mice were examined clinically during the

study and neurotoxic signs characterized by posterior paralysis were observed in 33 males after 59 weeks

and 3 females after 80 weeks in the 10 ppm group. A marked increase in mortality and a significant

decrease in body weight gain were also observed in the 10 ppm male dose group, beginning at 60 weeks. 

Post mortem examination revealed toxic encephalopathy consisting of neuronal necrosis of the brain and

toxic peripheral sensory neuropathy in both sexes of the 10 ppm group. An increased incidence of

chronic nephropathy was observed in the 2 and 10 ppm males. Based upon this study a NOAEL of 0.04

mg/kg-day and a LOAEL of 0.17 mg/kg-day was determined. These results indicated that B6C3F1 mice

are more sensitive to the neurotoxic effects of MeHg than ICR mice. 

KINETICS: MeHg in the diet is almost completely absorbed into the bloodstream. Animal studies

indicate (Walsh, 1982) that age has no effect on the efficiency of the gastrointestinal absorption, which is

usually in excess of 90%. From the bloodstream MeHg is distributed to all tissues, and distribution is

complete within 4 days in humans. The time necessary to reach peak brain levels from a single oral dose

is 1 or 2 days longer than other tissues and at this time the brain contains 6% of the total dose. Also at

this time the brain concentration is six times that of the blood.

Methylmercury is converted to inorganic Hg in various tissues at different rates in mammals. 

The fraction of total Hg present as Hg++ depends on the duration of exposure and the time after

cessation of exposure. The percentages of total Hg present as inorganic Hg++ in tissues of the Iraqi

population exposed for 2 months were: whole blood 7%, plasma 22%, breast milk 39% and urine 73%. 

Measurements in the hepatic tissue of patients that had died was 16-40% of Hg++. 

The fecal pathway accounts for 90% of the total elimination of Hg in mammals after exposure to

MeHg. Essentially all Hg in feces is in the inorganic form. The process of fecal elimination begins with

biliary excretion with extensive recycling of both MeHg and Hg++ complexed with glutathione. 

Inorganic Hg is poorly absorbed across the intestinal wall, but MeHg is readily reabsorbed such that a

secretion-resorption cycle is established. The intestinal microflora convert MeHg to inorganic Hg.

Whole body half-times determined in human volunteers averaged 70 days with a range of 52-93

days. Observations of blood half-times is 50 days with a range of 39-70 days. Lactating women have a

significantly shorter whole body half-time of 42 days compared with 79 days in nonlactating women. 

Selenium is known to bioconcentrate in fish and it is thought that simultaneous ingestion of

selenium may offer a protective effect for the toxicity of MeHg based upon its antioxidant properties. 

Selenium has been observed to correlate with Hg levels in blood (Granjean and Weihe, 1992). 
B-44

 I.A.5 CONFIDENCE IN THE ORAL RfD

Study -- Medium

Data Base -- Medium

RfD -- Medium

The benchmark dose approach allowed use of the entire dose-response assessment and the

calculation of a value that was consistent with the traditional NOAEL/LOAEL approach. In addition, the

results of laboratory studies with nonhuman primates support the quantitative estimate of the

NOAEL/LOAEL range of the benchmark dose that was indicated by the human studies. The reported

literature covers detailed studies of human exposures with quantitation of MeHg by analysis of

specimens from affected mother-fetus pairs. A strength of the Marsh study is the fact that the

quantitative data came directly from the affected population and quantitation is based on biological

specimens obtained from affected individuals. Unfortunately, a threshold was not easily defined and

extended application of modeling techniques were needed to define the lower end of the dose-response

curve. This may indicate high variability of response to MeHg in the human mother-fetal pairs or

misclassification in assigning pairs to the cohort. Recent concerns expressed in the research community

relate to the applicability of a dose-response estimate based on a grain-consuming population when the

actual application is likely to help characterize risk for fish-consuming segments of the population. 

Confidence in the supporting data base is medium. Confidence in the RfD is medium.

 I.A.6 EPA DOCUMENTATION

Source Document -- U.S. EPA, 1995

This IRIS summary is included in The Mercury Study Report to Congress, which was reviewed

by OHEA and EPA's Mercury Work Group in November 1994. An interagency review by scientists from

other federal agencies took place in January 1995. The report was also reviewed by a panel of

non-federal external scientists in January 1995 who met in a public meeting on January 25-26. All

reviewers comments have been carefully evaluated and considered in the revision and finalization of this

IRIS summary. A record of these comments is summarized in the IRIS documentation files.

Other EPA Documentation -- U.S. EPA, 1980, 1984, 1987, 1988

Agency Work Group Review -- 12/02/85, 03/25/92, 02/17/94, 08/04/94, 09/08/94, 

09/22/94, 10/13/94, 11/23/94 

Verification Date -- 11/23/94

 I.A.7 EPA CONTACTS (ORAL RfD)

Rita Schoeny / OHEA -- (513)569-7544

Bruce Mintz / OST -- (202)260-9569
B-45

REFERENCES

Aberg, B., L. Ekman, R. Falk, U. Greitz, G. Persson and J. Snihs. 1969. Metabolism of methyl mercury

(203Hg) compounds in man. Arch. Environ. Health. 19: 478-484. 

Allen B.C., R.J. Kavlock, C.A. Kimmel and E.M. Faustman. 1994a. Dose response assessment for

developmental toxicity. II. Comparison of generic benchmark dose estimates with NOAELS. Fund.

Appl. Toxicol. 23: 487-495.

Allen, B.C., R.J. Kavlock, C.A. Kimmel and E.M. Faustman. 1994b. Dose response assessment for

developmental toxicity. III. Statistical models. Fund. Appl. Toxicol. 23: 496-509. 

Best, C.H. and N.B. Taylor. 1961. A Text in Applied Physiology, 7th ed. The Williams and Wilkins

Co., Baltimore, MD. p. 19, 29.

Bornhausen, M., H.R. Musch and H. Greim. 1980. Operant behavior performance changes in rats after

prenatal methylmercury exposure. Toxicol. Appl. Pharmacol. 56: 305-310. 

Burbacher, T.M., C. Monnett, L.S. Grant and N.K. Mottet. 1984. Methylmercury exposure and

reproductive dysfunction in the nonhuman primate. Toxicol. Appl. Pharmacol. 75: 18-24. 

Cox, C., T.W. Clarkson, D.O. Marsh and G.G. Myers. 1989. Dose-response analysis of infants

prenatally exposed to methylmercury: An application of a single compartment model to single-strand hair

analysis. Environ. Res. 49: 318-332. 

Faustman, E.M., B.A. Allen, R.J. Kavlock and C.A. Kimmel. 1994. Dose-response assessment for

developmental toxicity. 1. Characterization of database and determination of no observed adverse effect

level. Fund. Appl. Toxicol. 23: 478-486. 

Grandjean, P. and P. Weihe. 1993. Neurobehavioral effects of intrauterine mercury exposure: Potential

sources of bias. Environ. Res. 61: 176-183.

Gunderson, V.M., K.S. Grant, J.F. Fagan and N.K. Mottet. 1986. The effect of low-level prenatal

methylmercury exposure on visual recognition memory of infant crab-eating macaques. Child Dev. 57:

1076-1083.

Kershaw, T.G., T.W. Clarkson and P.H. Dhahir. 1980. The relationship between blood levels and the

dose of methylmercury in man. Arch. Environ. Health. 35(1): 28-36. 

Kjellstrom, T., P. Kennedy, S. Wallis and C. Mantell. 1986. Physical and mental development of

children with prenatal exposure to mercury from fish. Stage 1: Preliminary test at age 4. National

Swedish Environmental Protection Board, Report 3080 (Solna, Sweden). 

Kjellstrom, T., P. Kennedy, S. Wallis amd C. Mantell. 1989. Physical and mental development of

children with prenatal exposure to mercury from fish. Stage 2: Interviews and psychological tests at age

6. National Swedish Environmental Protection Board, Report 3642 (Solna, Sweden). 
B-46

Marsh, D.O., T.W. Clarkson, C. Cox et al. 1987. Fetal methylmercury poisoning: Relationship between

concentration in single strands of maternal hair and child effects. Arch. Neurol. 44: 1017-1022. 

McKeown-Eyssen, G.E., J. Ruedy and A. Neims. 1983. Methylmercury exposure in northern Quebec. 

II: Neurologic finds in children. Am. J. Epidemiol. 118(4): 470-479. 

Miettinen, J.K., T. Rahola, T. Hattula, K. Rissanen and M. Tillander. 1971. Elimination of 203-Hg

methylmercury in man. Ann. Clin. Res. 3: 116-122. 

Mitsumori, K., M. Hiarano, H. Ueda, K. Maiata and Y. Shirasu. 1990. Chronic toxicity and

carcinogenicity of methylmercury chloride in B6C3F1 mice. Fund. Appl. Toxicol. 14: 179-190. 

 

Munro, I.C., E.A. Nera, S.M. Charbonneau, B. Junkins and Z. Zawidzka. 1980. Chronic toxicity of

methylmercury in the rat. J. Environ. Pathol. Toxicol. 3(5-6): 437-447. 

Phelps, R.W., T.W. Clarkson, T.G. Kershaw and B. Wheatley. 1980. Interrelationships of blood and

hair mercury concentrations in a North American population exposed to methylmercury. Arch. Environ.

Health. 35: 161-168. 

Rice, D.C. 1989. Delayed neurotoxicity in monkeys exposed developmentally to methylmercury. 

Neurotox. 10: 645-650. 

Marsh, D.O., T.W. Clarkson, C. Cox et al. 1987. Fetal methylmercury poisoning: Relationship between

concentration in single strands of maternal hair and child effects. Arch. Neurol. 44: 1017-1022. 

Seafood Safety. 1991. Committee on Evaluation of the Safety of Fishery Products, Chapter on

Methylmercury: FDA Risk Assessment and Current Regulations, National Academy Press, Washington,

DC. p. 196-221.

Sherlock, J.C., D.G. Lindsay, J. Hislop, W.H. Evans and T.R. Collier. 1982. Duplication diet study on

mercury intake by fish consumers in the United Kingdom. Arch. Environ. Health. 37(5): 271-278. 

Snyder, W.S., M.T. Cook, L.R. Karhausen et al. 1975. International Commission of Radiological

Protection. No. 23: Report of a Task Group on Reference Man. Pergamon Press, NY. 

Suzuki, T., T. Hongo, J. Yoshinaga et al. 1993. The hair-organ relationship in mercury concentration in

contemporary Japanese. Arch. Environ. Health. 48: 221-229. 

Tsubaki, T.K. and K. Irukayama. 1977. Minamata Disease: Methylmercury Poisoning in Minamata and

Niigata, Japan. Elsevier Science Publishers, New York. p. 143-253. 

U.S. EPA. 1980. Ambient Water Quality Criteria Document for Mercury. Prepared by the Office of

Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH

for the Office of Water Regulation and Standards, Washington, DC. EPA/440/5-80/058. NTIS PB

81-117699. 

U.S. EPA. 1984. Mercury Health Effects Update: Health Issue Assessment. Final Report. Prepared by

the Office of Health and Environmental Assessment, Environmental Criteria and Assessment Office,
B-47

Cincinnati, OH for the Office of Air Quality Planning and Standards, Research Triangle Park, NC. 

EPA/600/8-84/019F. NTIS PB81-85-123925.

U.S. EPA. 1987. Peer Review Workshop on Mercury Issues. Environmental Criteria and Assessment

Office, Cincinnati, OH. Summary report. October 26-27. 

U.S. EPA. 1988. Drinking Water Criteria Document for Inorganic Mercury. Prepared by the Office of

Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH

for the Office of Drinking Water, Washington, DC. EPA/600/X-84/178. NTIS PB89-192207.

U.S. EPA. 1995. Mercury Study Report to Congress. Office of Research and Development,

Washington, DC. External Review Draft. EPA/600/P-94/002Ab.

Walsh, C.T. 1982. The influence of age on the gastrointestinal absorption of mercuric chloride and

methylmercury chloride in the rat. Environ. Res. 27: 412-420. 

WHO (World Health Organization). 1990. Environmental Health Criteria 101: Methylmercury. 

Geneva. 
B-48

 II. CARCINOGENICITY ASSESSMENT FOR LIFETIME EXPOSURE

Substance Name -- Methylmercury

CASRN -- 22967-92-6

Preparation Date -- 5/24/94

 II.A EVIDENCE FOR CLASSIFICATION AS TO HUMAN CARCINOGENICITY

 II.A.1 WEIGHT-OF-EVIDENCE CLASSIFICATION

Classification -- C; possible human carcinogen

Basis -- Based on inadequate data in humans and limited evidence of carcinogenicity in animals. Male

ICR and B6C3F1 mice exposed to methylmercuric chloride in the diet had an increased incidence of

renal adenomas, adenocarcinomas and carcinomas. The tumors were observed at a single site and in a

single species and single sex. The renal epithelial cell hyperplasia and tumors were observed only in the

presence of profound nephrotoxicity and were suggested to be a consequence of reparative changes in the

cells. Several nonpositive cancer bioassays were also reported. Although genotoxicity test data suggest

that methylmercury is capable of producing chromosomal and nuclear damage, there are also nonpositive

genotoxicity data. 

 II.A.2 HUMAN CARCINOGENICITY DATA

Inadequate. Three studies were identified that examined the relationship between methylmercury

exposure and cancer. No persuasive evidence of increased carcinogenicity attributable to methylmercury

exposure was observed in any of the studies. Interpretation of these studies, however, was limited by

poor study design and incomplete descriptions of methodology and/or results. 

Tamashiro et al. (1984) evaluated the causes of death in 334 subjects from the Kumamoto

Prefecture who had been diagnosed with Minamata disease (methylmercury poisoning) and died between

1970 and 1980. The subjects involved fishermen and their families who had been diagnosed with the

disease; thus, Minamata disease was used as a surrogate for methylmercury exposure. The controls were

selected from all deaths that had occurred in the same city or town as the cases and were matched on the

basis of sex, age at death (within 3 years) and year of death; two controls were matched to each subject. 

Malignant neoplasms were designated as the underlying cause of death in 14.7% (49/334) of the subjects

and 20.1% (134/668) of the controls. For 47 subjects in which Minamata disease was listed as the

underlying cause of death, the investigators reanalyzed the mortality data and selected one of the

secondary causes to be the underlying cause of death in order to allow examination of the subjects and

controls under similar conditions and parameters. The three subjects for which Minamata disease was

listed as the only cause of death were excluded from further analysis. Using the Mantel-Haenzel method

to estimate odds ratios, no significant differences were observed between the subjects and controls with

respect to the proportion of deaths due to malignant neoplasms among males, females or both sexes

combined. The estimated odds ratios and 95% confidence intervals were 0.84 (0.49-1.43), 0.58

(0.28-1.21) and 0.75 (0.50-1.11) for males, females and both sexes combined. Similarly, no increases in

odds ratio were observed among the subjects relative to the controls when malignant neoplasms were
B-49

identified as a secondary cause of death or were listed on death certificates as one of the multiple causes

of death. These data suggest that cancer incidence was not increased in persons with overt signs of

methylmercury poisoning when compared with persons for whom no diagnosis of methylmercury

poisoning had been made. Interpretation is limited by potential bias in designating the cause of death

among patients with known Minamata disease and by the uncertainty regarding the extent of

methylmercury exposure and undiagnosed Minamata disease among the controls. In a subsequent study,

Tamashiro et al. (1986) compared the mortality patterns (between 1970 and 1981) among residents of the

Fukuro and Tsukinoura districts in the Kumamoto Prefecture (inhabited mainly by fishermen and their

families) with that of age-matched residents of Minamata City (also in the Kumamoto Prefecture) who

died between 1972 and 1978. In this study, high exposure to methylmercury was inferred from residence

in a district believed to have higher intake of local seafood. By contrast, in the 1984 study described

above, high methylmercury exposure was inferred from a diagnosis of Minamata disease. A total of 416

deaths were recorded in the Fukuro and Tsukinoura districts in 1970-1981, and 2325 deaths were

recorded in Minamata City in 1972-1978. No statistically significant increase in the overall cancer

mortality rate was observed; however, an increase in the mortality rate due to liver cancer was observed

(SMR, 207.3; 95% CI, 116.0-341.9). Analysis of mortality by sex showed a statistically significant

increase in the rate of liver cancer only among males (SMR, 250.5; 95% CI, 133.4-428.4). Males also

had statistically significant higher mortality due to chronic liver disease 

and cirrhosis. The authors note that these results should be interpreted with caution because the

population of the Fukuro and Tsukinoura districts had higher alcohol consumption and a higher

prevalence of hepatitis B (a predisposing factor for hepatocellular cancer). Interpretation of these results

is also limited by an incomplete description of the methodology used to calculate the SMRs; it is unclear

whether the study authors used appropriate methods to compare mortality data collected over disparate

time frames (12 years for exposed and 7 years for controls). 

In a study from Poland, Janicki et al. (1987) reported a statistically significant increase in

mercury content of hair in leukemia patients (0.92 +/-1.44 ppm [sic]; n=47) relative to that in healthy

unrelated patients (0.49 +/-0.41 ppm; n=79). Similarly, the mercury content in the hair of a subgroup of

leukemia patients (0.69 +/- 0.75; n=19) was significantly greater than that in healthy relatives who had

shared the same residence for at least 3 years preceding the onset of the disease (0.43 +/- 0.24 ppm;

n=52). When patients with specific types of leukemia were compared with the healthy unrelated subjects

(0.49 +/- 0.41 ppm; n=79), only those with acute leukemia (type not specified; 1.24 +/- 1.93 ppm; n=23)

had a significantly increased hair mercury content. No significant differences in hair mercury content

were observed in 9 patients with chronic granulocytic leukemia or 15 patients with chronic lymphocytic

leukemia when compared with the unrelated, healthy controls. The authors inferred that acute leukemia

was associated with increased level of mercury in hair. This study is of limited use for cancer risk

assessment because of the following: uncertainty regarding the correlation between the chronology of

incorporation of mercury in the hair and onset of the disease; the small population studied; the failure to

describe adequately the characteristics of the leukemia patients or healthy controls (age 

distribution, length of residence in the region, criteria for inclusion in the study); uncertainty regarding

the source of mercury exposure (the authors presumed that exposure was the result of use of

methylmercury-containing fungicides); and the failure to address exposure to other chemicals or adjust

for other leukemia risk factors. Furthermore, the variability of hair mercury 

content was large, and the mean hair mercury levels were within normal limits for all groups. Thus, the

statistical significance may have been due to chance. 

The carcinogenic effects of organomercury seed dressing exposure were investigated in a series

of case-control studies for incidence of soft-tissue sarcomas (Eriksson et al., 1981, 1990; Hardell and
B-50

Eriksson, 1988) or malignant lymphomas (Hardell et al., 1981). These studies were conducted in

Swedish populations exposed to phenoxyacetic acid herbicides or chlorophenols (the exposures of

primary interest in the studies), organomercury seed dressings, or other pesticides. Exposure frequencies

were derived from questionnaires and/or interviews. Control groups from the same region of the country

were matched to cases based on vital status. A total of 402 cases of soft-tissue sarcoma and (among

persons not exposed to phenoxyacetic acid herbicides) 128 cases of malignant lymphoma were reported. 

In each study, the odds ratio for exposure to organomercury in seed dressings and the incidence of

sarcoma or lymphoma was either <1.0 or the range of the 95% confidence interval for the odds ratio

included 1.0; therefore, no association was indicated for organomercury exposure and soft-tissue sarcoma

or malignant lymphoma. The study subjects were likely to have experienced exposures to the other

pesticides and chemicals. 

 II.A.3 ANIMAL CARCINOGENICITY DATA

Limited. Three dietary studies in two strains of mice indicate that methylmercury is

carcinogenic. Interpretation of two of the positive studies was complicated by observation of tumors

only at doses that exceeded the MTD. A fourth dietary study in mice and four dietary studies in rats

failed to indicate carcinogenicity associated with methylmercury exposure. Interpretation of four of the

nonpositive studies was limited because of deficiencies in study design or failure to achieve an MTD.

Methylmercuric chloride (>99% pure) was administered in the diet at levels of 0, 0.4, 2 or 10

ppm (0, 0.03, 0.15 and 0.73 (mg/kg)/day in males and 0, 0.02, 0.11 and 0.6 (mg/kg)/day in females) to 60

ICR mice/sex/group for 104 weeks (Hirano et al., 1986). Interim sacrifices (6/sex/group) were conducted

at 26, 52 and 78 weeks. Complete histopathological examinations were performed on all animals found

dead, killed in extremis or killed by design. Mortality, group mean body weights and food consumption

were comparable to controls. The first renal tumor was observed at 58 weeks in a high-dose male, and

the incidence of renal epithelial tumors (adenomas or adenocarcinomas) was significantly increased in

high-dose males (1/32, 0/25, 0/29 and 13/26 in the control, low-, mid- and high-dose groups,

respectively). Ten of the 13 tumors in high-dose males were adenocarcinomas. These tumors were

described as solid type or cystic papillary types of adenocarcinomas. No invading proliferation into the

surrounding tissues was observed. The incidence of renal epithelial adenomas was not significantly

increased in males, and no renal adenomas or adenocarcinomas were observed in any females studied. 

Focal hyperplasia of the tubular epithelium was reported to be increased in high-dose males (13/59; other

incidences not reported). Increases in non-neoplastic lesions in high-dose animals provided evidence that

an MTD was exceeded. Non-neoplastic lesions reported as increased in treated males included the

following: epithelial degeneration of the renal proximal tubules; cystic kidney; urinary cast and pelvic

dilatation; and decreased spermatogenesis. Epithelial degeneration of the renal proximal tubules and

degeneration or fibrosis of the sciatic nerve was reported in high-dose females.

Methylmercuric chloride (>99% pure) was administered in the diet at levels of 0, 0.4, 2 or 10

ppm (0, 0.3, 0.14 and 0.69 (mg/kg)/day in males and 0, 0.03, 0.13 and 0.60 (mg/kg)/day in females) to 60

B6C3F1 mice/sex/group for 104 weeks (Mitsumori et al., 1990). In high-dose males, a marked increase

in mortality was observed after week 60 (data presented graphically; statistical analyses not performed by

authors). Survival at study termination was approximately 50, 60, 60 and 20% in control, low-, mid- and

high-dose males, respectively, and 58, 68, 60 and 60% in control, low-, mid- and high-dose females,

respectively. The cause of the high mortality was not reported. At study termination, the mean body
B-51

weight in high-dose males was approximately 67% of controls and in high-dose females was

approximately 90% of controls (data presented graphically; statistical analyses not performed by study

authors). The incidence of focal hyperplasia of the renal tubules was significantly increased in high-dose

males (14/60; the incidence was 0/60 in all other groups). The incidence of renal epithelial carcinomas

(classified as solid or cystic papillary type) was also significantly increased in high-dose males (13/60;

the incidence was 0/60 in all other groups). The incidence of renal adenomas (classified as solid or

tubular type) was also significantly increased in high-dose males; the incidence was 0/60, 0/60, 1/60 and

5/60 in control, low-, mid- and high-dose males, respectively, and 0/60, 0/60, 0/60 and 1/60 in control

low-, mid- and high-dose females, respectively. No metastases were seen in the animals. The incidences

of a variety of non-neoplastic lesions were increased in the high-dose mice including the following: 

sensory neuropathy; neuronal necrosis in the cerebrum; neuronal degeneration in the cerebellum; and

chronic nephropathy of the kidney. Males exhibited tubular atrophy of the testis (1/60, 5/60, 2/60 and

54/60 in control, low-, mid- and high-dose, respectively) and ulceration of the glandular stomach (1/60,

1/60, 0/60 and 7/60 in control, low-, mid- and high-dose males, respectively). An MTD was achieved in

mid-dose males and high-dose females. High mortality in high-dose males indicated that the MTD was

exceeded in this group. 

Mitsumori et al. (1981) administered 0, 15 or 30 ppm of methylmercuric chloride (>99% pure) in

the diet (0, 1.6 and 3.1 (mg/kg)/day) to 60 ICR mice/sex/group for 78 weeks. Interim sacrifices of up to

6/sex/group were conducted at weeks 26 and 52. Kidneys were microscopically examined from all

animals that died or became moribund after week 53 or were killed at study termination. Lungs from

mice with renal masses and renal lymph nodes showing gross abnormalities were also examined. 

Survival was decreased in a dose-related manner; at week 78 survival was 40, 10 and 0% in control, lowand high-dose males, respectively, and 55, 30 and 0%, in control, low- and high-dose females,

respectively (statistical analyses not performed). The majority of high-dose mice (85% males and 98%

females) died by week 26 of the study. Examination of the kidneys of mice that died or were sacrificed

after 53 weeks showed a significant increase in renal tumors in low-dose males (13/16 versus 1/37 in

controls). The incidence of renal epithelial adenocarcinomas in control and low-dose males was 0/37 and

11/16, respectively. The incidence of renal epithelial adenomas in control and low-dose males was 1/37

and 5/16, respectively. No renal tumors were observed in females in any group. No metastases to the

lung or renal lymph nodes were observed. Evidence of neurotoxicity and renal pathology were observed

in the treated mice at both dose levels. The high mortality in both groups of treated males and in

high-dose females indicated that the MTD was exceeded in these groups. (Note: Hirano et al. (1986)

was a followup to this study.)

Mitsumori et al. (1983, 1984) administered diets containing 0, 0.4, 2 or 10 ppm of

methylmercuric chloride (0, 0.011, 0.05 and 0.28 (mg/kg)/day in males; 0, 0.014, 0.064 and 0.34

(mg/kg)/day in females) to 56/sex/group Sprague-Dawley rats for up to 130 weeks. Interim sacrifices of

10/group (either sex) were conducted at weeks 13 and 26 and of 6/group (either sex) at weeks 52 and 78. 

Mortality was increased in high-dose males and females. At week 104, survival was approximately 55,

45, 75 and 10% in control, low-, mid-and high-dose males, respectively, and 70, 75, 75 and 30% in

control, low-, mid- and high-dose females, respectively (data presented graphically). All males in the

high-dose group had died by week 119. Body weight gain was significantly decreased in high-dose

males starting after week 44 and females 

after 44 weeks (approximately 10-20%, data presented graphically). No increase in tumor incidence was

observed in either males or females. Noncarcinogenic lesions that were significantly increased in

high-dose rats included the following: degeneration in peripheral nerves and the spinal cord (both

sexes); degeneration of the proximal tubular epithelium (both sexes); severe chronic nephropathy
B-52

(females); parathyroid hyperplasia (both sexes); polyarteritis nodosa and calcification of arterial wall

(females); fibrosis of bone (females); bile duct hyperplasia (males); and hemosiderosis and

extramedullary hematopoiesis in the spleen (males). Mid-dose males exhibited significantly increased

degeneration of the proximal tubular epithelium and hyperplasia of the parathyroid. An MTD was

achieved in mid-dose males and exceeded in high-dose males and high-dose females.

No tumor data were reported in a study using Wistar rats (Munro et al., 1980). Groups of 50

Wistar rats/sex/dose were fed diets containing methylmercury; doses of 2, 10, 50 and 250 (ug/kg)/day

were fed for 26 months. High-dose female rats exhibited reduced body weight gains and showed minimal

clinical signs of neurotoxicity; however, high-dose male rats showed overt clinical signs of neurotoxicity,

decreased hemoglobin and hematocrit values, reduced weight gains and significantly increased mortality. 

Histopathologic examination of the high-dose rats of both sexes revealed demyelination of dorsal nerve

roots and peripheral nerves. Males showed severe dose-related kidney damage, and females had minimal

renal damage.

No increase in tumor incidence or decrease in tumor latency was observed in another study using

rats of an unspecified strain (Verschuuren et al., 1976). Groups of 25 female and 25 male rats were

administered methylmercuric chloride at dietary levels of 0, 0.1, 0.5 and 2.5 ppm (0, 0.004, 0.02 and 0.1

(mg/kg)/day) for 2 years. No significant effects were observed on growth or food intake except for a 6%

decrease (statistically significant) in body weight gain at 60 weeks in high-dose females. Survival was

72, 68, 48 and 48% in control, low-, mid- and high-dose males, respectively, and 76, 60, 64 and 56% in

control, low-, mid- and high-dose females, respectively (statistical significance not reported). Increases

in relative kidney weights were observed in both males and females at the highest dose. No effects on

the nature or incidence of pathological lesions were observed, and tumors were reported to have been

observed with comparable incidence and latency among all of the groups. This study was limited by the

small sample size. 

No increase in tumor incidence was observed in a study using white Swiss mice (Schroeder and

Mitchener, 1975). Groups of mice (54/sex/group) were exposed until death to methylmercuric acetate in

the drinking water at two doses. The low-dose group received 1 ppm methylmercuric acetate (0.19

(mg/kg)/day). The high-dose group received 5 ppm methylmercuric acetate (0.95 (mg/kg)/day) for the

first 70 days and then 1 ppm thereafter, due to high mortality (21/54 males and 23/54 females died prior

to the dose reduction). Survival among the remaining mice was not significantly different from controls. 

Significant reductions in body weight were reported in high-dose males (9-15% lower than controls) and

high-dose females (15-22% lower than controls) between 2 and 6 months of age. After dying, mice were

weighed and dissected; gross tumors were counted, and limited histopathologic sections were made of

heart, lung, liver, kidney and spleen for microscopic examination. This study is limited because

complete histological examinations were not performed.

No increase in tumor incidence was observed in a multiple-generation reproduction study using

Sprague-Dawley rats (Newberne et al., 1972). Groups of rats (30/sex) were given semisynthetic diets

supplemented with either casein or a fish protein concentrate to yield dietary levels of 0.2 ppm

methylmercury (0.008 (mg/kg)/day). Another group of controls received untreated rat chow. Rats that

received diets containing methylmercury during the 2-year study had body weights and hematology

comparable to controls. Detailed histopathological analyses revealed no lesions of the brain, liver, or

kidney that were attributable to methylmercury exposure. Mortality data were not presented. 

Interpretation of these data is limited by the somewhat small group sizes and failure to achieve an MTD.
B-53

 II.A.4 SUPPORTING DATA FOR CARCINOGENICITY

Blakley (1984) administered methylmercuric chloride to female Swiss mice (number/group not

specified) in drinking water at concentrations of 0, 0.2, 0.5 or 2.0 mg/L for 15 weeks (approximately 0,

0.03, 0.07 and 0.27 (mg Hg/kg)/day). At the end of week 3, a single dose of 1.5 mg/kg of urethane was

administered intraperitoneally to 16-20 mice/group. No effects on weight gain or food consumption were

observed. Lung tumor incidence in mice not administered urethane (number/group not specified) was

less than one tumor/mouse in all groups. Statistically significant trends for increases in the number and

size of lung adenomas/mouse with increasing methylmercury dose were observed; the number of

tumors/mouse was 21.5, 19.4, 19.4 and 33.1 in control, low-, mid- and high-dose mice, respectively, and

the tumor size/mouse was 0.70, 0.73, 0.76 and 0.76 mm in control, low-, mid- and high-dose mice,

respectively. The study authors suggest that the increase in tumor number and size may have been

related to the immunosuppressive activity of methylmercury. It should be noted that this study is

considered a short-term bioassay, and pulmonary adenomas were the only tumor type evaluated.

Humans ingesting methylmercury-contaminated foods have been reported to experience

chromosomal aberrations (Skerfving et al., 1970, 1974) or SCE (Wulf et al., 1986); however,

interpretation of these studies is limited by methodological deficiencies.

As reviewed in WHO (1990), methylmercury is not a potent mutagen but appears to be capable

of causing chromosome damage and nuclear perturbations in a variety of systems. In Bacillus subtilis,

methylmercury produced DNA damage (Kanematsu et al., 1980). Methylmercury produced

chromosomal aberrations and aneuploidy in human peripheral lymphocytes (Betti et al., 1992), SCE in

human lymphocytes (Morimoto et al., 1982), and DNA damage in human nerve and lung cells as well as

Chinese hamster V-79 cells and rat glioblastoma cells (Costa et al., 1991). 

Bone marrow cells of cats treated with methylmercury in a study by Charbonneau et al. (1976)

were examined by Miller et al. (1979). The methylmercury treatment resulted in an increased number of

nuclear abnormalities and an inhibition of DNA repair capacity. Methylmercury induced a weak

mutagenic response in Chinese hamster V-79 cells (Fiskesjo, 1979). Methylmercury also induced

histone protein perturbations and influenced factors regulating nucleolus-organizing activity (WHO,

1990). Moreover, methylmercury has been reported to interfere with gene expression in cultures of

glioma cells (WHO, 1990). Mailhes (1983) reported a significant increase in the number of hyperploid

oocysts in Lak:LVG Syrian hamsters fed methylmercury; however, no evidence of chromosomal damage

was reported. Suter (1975) concluded that strain-specific differences exist with respect to the ability of

methylmercury to produce dominant lethal effects in mice. Nondisjunction and sex-linked recessive

lethal mutations were observed in Drosophila melanogaster treated with methylmercury (Ramel, 1972 as

cited in U.S. EPA, 1985). Methylmercury produced single strand breaks in DNA in cultured L5178Y

cells (Nakazawa et al., 1975). 

Negative studies have also been reported. Methylmercury acetate was reported to be negative in

a Salmonella typhimurium assay and a mouse micronucleus assay (Heddle and Bruce, 1977, as reported

in Jenssen and Ramel, 1980). Methylmercury was not mutagenic and did not cause recombination in

Saccharomyces cerevisiae but did slightly increase chromosomal nondisjunction (Nakai and Machida,

1973). Matsumoto and Spindle (1982) reported no significant increase in SCE in developing mouse

embryos; they did report, however, that the developing mouse embryos were highly sensitive to in vitro

treatment with methylmercury. 
B-54

 II.B QUANTITATIVE ESTIMATE OF CARCINOGENIC RISK FROM ORAL EXPOSURE

None. The two studies by Mitsumori et al. (1981, 1990) were limited by high mortality in the

high-dose males, the only group to exhibit a statistically significant increase in tumor incidence. Tumors

were observed only in those dose groups in which the MTD had been exceeded. The study by Hirano et

al. (1986) was not limited by low survival, but the tumors were observed in conjunction with

nephrotoxicity and, thus, their incidence may have been a high-dose phenomenon that would not be

expected to occur at low doses. The tumors appeared to originate from focal hyperplasia of the tubular

epithelium induced as a reparative change. The hyperplasia was not observed in tubular epithelium that

was undergoing early degenerative changes. Thus, the tumors may not occur where degenerative

changes do not occur. The genotoxicity data indicate that methylmercury is not a potent mutagen but

may produce chromosomal damage; these data do not support a hypothesis that methylmercury is a

genotoxic carcinogen. It appears, rather, that methylmercury exerts its carcinogenic effect only at high

dose, at or above an MTD. Because the linearized multistage procedure is based on the assumption of

linearity at low doses, the relevance of deriving a slope factor based on data for which a threshold may

exist is questionable.

It is likely that systemic non-cancer effects would be seen at methylmercury exposures lower

than those required for tumor formation. Long-term administration of methylmercury to experimental

animals produces overt symptoms of neurotoxicity at daily doses an order of magnitude lower than those

required to induce tumors in mice.

 II.C QUANTITATIVE ESTIMATE OF CARCINOGENIC RISK FROM INHALATION

EXPOSURE

None.

 II.D EPA DOCUMENTATION, REVIEW, AND CONTACTS (CARCINOGENICITY

ASSESSMENT)

 II.D.1 EPA DOCUMENTATION

Source Documents -- U.S. EPA, 1995

This IRIS summary is included in The Mercury Study Report to Congress which was reviewed

by OHEA and EPA's Mercury Work Group in November 1994. An interagency review by scientists from

other federal agencies took place in January 1995. The report was also reviewed by a panel of

non-federal external scientists in January 1995 who met in a public meeting on January 25-26. All

reviewers comments have been carefully evaluated and considered in the revision and finalization of this

IRIS summary. A record of these comments is summarized in the IRIS documentation files. 
B-55

 II.D.2 REVIEW (CARCINOGENICITY ASSESSMENT)

Agency Work Group Review -- 03/03/94

Verification Date -- 03/03/94

 II.D.3 U.S. EPA CONTACTS (CARCINOGENICITY ASSESSMENT)

Rita Schoeny / OHEA -- (513)569-7544

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mercury in human lymphocytes. Mutat. Res. 281(4): 255 260. 

Blakley, B.R. 1984. Enhancement of urethane-induced adenoma formation in Swiss mice exposed to

methylmercury. Can. J. Comp. Med. 48: 299 302. 

Charbonneau, S.M., I.C. Munro, E.A. Nera et al. 1976. Chronic toxicity of methylmercury in the adult

cat. Toxicology. 5: 337-349. 

Costa, M., N.T. Christie, O. Cantoni, J.T. Zelikoff, X.W. Wang and T.G. Rossman. 1991. DNA damage

by mercury compounds: An overview. In: Advances in Mercury Toxicity, T. Suzuki, N. Imura and

T.W. Clarkson, Ed. Plenum Press, New York, NY. p. 255-273. 

Eriksson, M., L. Hardell, N.O. Berg, T. Moller and O. Axelson. 1981. Soft-tissue sarcomas and

exposure to chemical substances: A case-referent study. Br. J. Ind. Med. 38: 27 33. 

Eriksson, M., L. Hardell and H.-O. Adami. 1990. Exposure to dioxins as a risk factor for soft-tissue

sarcoma: A population-based case-control study. J. Natl. Cancer Inst. 82(6): 486 490. 

Fiskesjo, G. 1979. Two organic mercury compounds tested for mutagenicity in mammalian cells by use

of the cell line V 79-4. Hereditas. 90: 103-109.

Hardell, L. and M. Eriksson. 1988. The association between soft-tissue sarcomas and exposure to

phenoxyacetic acids. A new case-referent study. Cancer. 62: 652 656. 

Hardell, L., M. Eriksson, P. Lenner and E. Lundgren. 1981. Malignant lymphoma and exposure to

chemicals, especially organic solvents, chlorophenols and phenoxy acids: A case-control study. Br. J.

Cancer. 43: 169 176. 

Heddle, J.R. and W.R. Bruce. 1977. Comparison of the micronucleus and sperm assay for mutagenicity

with the carcinogenic activities of 61 different agents. In: Origins of Human Cancer, H.H. Hiatt, J.D.

Watson, J.A. Winsten, Ed. Vol. 4. Cold Spring Harbor Conferences. 
B-56

Hirano, M., K. Mitsumori, K. Maita and Y. Shirasu. 1986. Further carcinogenicity study on

methylmercury chloride in ICR mice. Jpn. J. Vet. Sci. 48(1): 127 135. 

Janicki, K., J. Dobrowolski and K. Krasnicki. 1987. Correlation between contamination of the rural

environment with mercury and occurrence of leukemia in men and cattle. Chemosphere. 16: 253 257. 

Jenssen, D. and C. Ramel. 1980. The micronucleus test as part of a short-term mutagenicity test

program for the prediction of carcinogenicity evaluated by 143 agents tested. Mutat. Res. 75: 191 202. 

Kanematsu, N., M. Hara and T. Kada. 1980. Rec assay and mutagenicity studies on metal compounds. 

Mutat. Res. 77: 109-116. 

Mailhes, J.B. 1983. Methylmercury effects on Syrian hamster metaphase II oocyte chromosomes. 

Environ. Mutagen. 5: 679-686. 

Matsumoto, N. and A. Spindle. 1982. Sensitivity of early mouse embryos to methylmercury toxicity. 

Toxicol. Appl. Pharmacol. 64: 108-117. 

Miller, C.T., Z. Zawidska, E. Nagy and S.M. Charbonneau. 1979. Indicators of genetic toxicity in

leukocytes and granulocytic precursors after chronic methylmercury ingestion by cats. Bull. Environ.

Contam. Toxicol. 21:296-303. 

Mitsumori, K., K. Maita, T. Saito, S. Tsuda and Y. Shirasu. 1981. Carcinogenicity of methylmercury

chloride in ICR mice: Preliminary note on renal carcinogenesis. Cancer Lett. 12: 305 310. 

Mitsumori, K., K. Takahashi, O. Matano, S. Goto and Y. Shirasu. 1983. Chronic toxicity of

methylmercury chloride in rats: Clinical study and chemical analysis. Jpn. J. Vet. Sci. 45(6): 747-757.

Mitsumori, K., K. Maita and Y. Shirasu. 1984. Chronic toxicity of methylmercury chloride in rats: 

Pathological study. Jpn. J. Vet. Sci. 46(4): 549-557. 

Mitsumori, K., M. Hirano, H. Ueda, K. Maita and Y. Shirasu. 1990. Chronic toxicity and

carcinogenicity of methylmercury chloride in B6C3F1 mice. Fund. Appl. Toxicol. 14: 179 190. 

Morimoto, K., S. Iijima and A. Koizumi. 1982. Selenite prevents the induction of sister-chromatid

exchanges by methyl mercury and mercuric chloride in human whole-blood cultures. Mutat. Res. 102:

183-192. 

Munro, I., E. Nera, S. Charbonneau, B. Junkins and Z. Zawidzka. 1980. Chronic toxicity of

methylmercury in the rat. J. Environ. Pathol. Toxicol. 3: 437-447. 

Nakai, S. and I. Machida. 1973. Genetic effect of organic mercury on yeast. Mutat. Res. 21(6): 348. 

Nakazawa, N., F. Makino and S. Okada. 1975. Acute effects of mercuric compounds on cultured

mammalian cells. Biochem. Pharmacol. 24: 489-493.

Newberne, P.M., O. Glaser and L. Friedman. 1972. Chronic exposure of rats to methyl mercury in fish

protein. Nature. 237: 40-41. 
B-57

Ramel, C. 1972. Genetic effects. In: Mercury in the Environment - An Epidemiological and

Toxicological Appraisal, L. Friberg and J. Vostal, Ed. CRC Press, Cleveland, OH. p. 169-181. (Cited in

U.S. EPA, 1985).

Schroeder, H. and M. Mitchener. 1975. Life-time effects of mercury, methylmercury, and nine other

trace metals in mice. J. Nutr. 105: 452 458.

Skerfving, S., K. Hansson and J. Lindsten. 1970. Chromosome breakage in humans exposed to methyl

mercury through fish consumption. Arch. Environ. Health. 21: 133-139. 

Skerfving, S., K. Hansson, C. Mangs, J. Lindsten and N. Ryman. 1974. Methylmercury-induced

chromosome damage in man. Environ. Res. 7: 83-98. 

Suter, K.E. 1975. Studies on the dominant-lethal and fertility effects of the heavy metal compounds

methylmercuric hydroxide, mercuric chloride, and cadmium chloride in male and female mice. Mutat.

Res. 30: 365-374. 

Tamashiro, H., M. Arakaki, H. Akagi, M. Futatsuka and L.H. Roht. 1984. Causes of death in Minamata

disease: Analysis of death certificates. Int. Arch. Occup. Environ. Health. 54: 135-146. 

Tamashiro H., Arakaki M., Futatsuka M. and E.S. Lee. 1986. Methylmercury exposure and mortality in

southern Japan: A close look at causes of death. J. Epidemiol. Comm. Health. 40: 181-185. 

U.S. EPA. 1980. Ambient Water Quality Criteria Document for Mercury. Prepared by the Office of

Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH

for the Office of Water Regulation and Standards, Washington, DC. EPA/440/5-80/058. NTIS

PB81-117699. 

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the Office of Health and Environmental Assessment, Environmental Criteria and Assessment Office,

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EPA/600/8-84/019F. NTIS PB81-85-123925. 

U.S. EPA. 1984b. Health Effects Assessment for Mercury. Prepared by the Office of Health and

Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH for the

Office of Emergency and Remedial Response, Washington, DC. EPA/540/1086/042. NTIS

PB86-134533/AS. 

U.S. EPA. 1988. Drinking Water Criteria Document for Inorganic Mercury. Prepared by the Office of

Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH

for the Office of Drinking Water, Washington, DC. EPA/600/X-84/178. NTIS PB89-192207. 

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Assessment (Draft). Prepared by the Office of Health and Environmental Assessment, Environmental

Criteria and Assessment Office, Cincinnati, OH for the Office of Air Quality Planning and Standards,

Research Triangle Park, NC. EPA/600/R-92/199. 
B-58

U.S. EPA. 1995. Mercury Study Report to Congress. Office of Research and Development,

Washington, DC. External Review Draft. EPA/600/P-94/002Ab. 

Verschuuren, H.G., R. Kroes, E.M. Den Tonkelaar et al. 1976. Toxicity of methylmercury chloride in

rats. III. Long-term toxicity study. Toxicology. 6: 107 123. 

WHO (World Health Organization). 1990. Methyl mercury. Vol. 101. Geneva, Switzerland: World

Health Organization, Distribution and Sales Service, International Programme on Chemical Safety. 

Wulf, H.C., N. Kromann, N. Kousgaard, J.C. Hansen, E. Niebuhr and K. Alboge. 1986. Sister

chromatid exchange (SCE) in Greenlandic eskimos: Dose-response relationship between SCE and seal

diet, smoking, and blood cadmium and mercury concentrations. Sci. Total Environ. 48: 81-94.
APPENDIX C

ATTENDEES OF U.S. EPA PEER REVIEW WORKSHOP

ON MERCURY ISSUES
C-1

LIST OF WORKSHOP ATTENDEES

Giuseppe Andres Ernest Foulkes

State University of New York University of Cincinnati College of Medicine

206 Farber Hall Department of Environmental Health

Main Street Campus Cincinnati, OH 45267-0056

Department of Pathology 513-872-5769

Buffalo, NY 14214

716-831-2846 Kris Khanna

Karen Blackburn U.S. EPA

Environmental Criteria and Assessment Office 401 M Street, S.W.

U.S. EPA Washington, DC 20460

26 West Martin Luther King Drive 202-382-7588

Cincinnati, OH 45268

513-569-7569 Loren D. Koller

Harlal Choudhury Medicine

Environmental Criteria and Assessment Office Corvallis, OR 97331

U.S. EPA 503-754-2098

26 West Martin Luther King Drive

Cincinnati, OH 45268 W. Bruce Peirano

513-569-7536 Environmental Criteria and Assessment Office

Tom Clarkson 26 West Martin Luther King Drive

University of Rochester Cincinnati, OH 45268

P.O. Box EHSC 513-569-7540

University of Rochester Medical School

Rochester, NY 14642 David Reisman

716-275-3911 Environmental Criteria and Assessment Office

Michael Dieter 26 West Martin Luther King Drive

NTP, NIEHS, NIH Cincinnati, OH 45268

P.O. Box 12233 513-569-7588

Research Triangle Park, NC 27709

919-541-3368 Paul L. Richter

Michael Dourson Environmental Protection

Environmental Criteria and Assessment Office CN413

U.S. EPA 401 East State Street - Sixth Floor

26 West Martin Luther King Drive Trenton, NJ 08625

Cincinnati, OH 45268 609-984-9759

513-569-7544

Office of Drinking Water

Oregon State University College of Veterinary

U.S. EPA

U.S. EPA

New Jersey Department of
C-2

Linda Saunders Cindy Sonich-Mullin

Eastern Research Group, Inc. Environmental Criteria and Assessment Office

6 Whittemore Street U.S. EPA

Arlington, MA 02174 26 West Martin Luther King Drive

617-648-7800 Cincinnati, OH 45268

Heidi Schultz

Eastern Research Group, Inc. Bob Vanderslice

6 Whittemore Street Office of Drinking Water

Arlington, MA 02174 U.S. EPA

617-648-7800 401 M Street, S.W.

513-569-7523

Washington, DC 20460

202-475-6711
APPENDIX D

HEALTH EFFECTS OF MERCURY AND MERCURY COMPOUNDS

UNCERTAINTY ANALYSIS OF THE METHYLMERCURY RfD
D-1

D.1 Introduction and Background

The purpose of the analysis in this appendix is two-fold: first, to determine plausible bounds on

uncertainty associated with the data and dose conversions used to derive the methylmercury Reference

Dose (RfD); second, to compare the RfD to estimated distributions of human population thresholds for

adverse effects. The analysis presented in this appendix is a modeled estimate of the human threshold for

specific health effects attributable to methylmercury exposure. The basis for the analysis and the RfD is

the data from the 1971 Iraqi methylmercury poisoning incident, specifically the data from the Marsh et

al. (1987) study. The analysis also includes studies pertinent to the conversion of mercury

concentrations in hair to estimated ingestion levels. The population studied in Marsh et al. (1987) is

hereafter referred to as the Iraqi cohort. The methylmercury RfD was based on a benchmark dose

calculated from the combined developmental effects of late walking, late talking, mental effects, seizures

and neurological effects (scores greater than 3 on a test) in children of women exposed during pregnancy;

benchmark doses for the individual developmental effects and for adult paresthesia were also calculated. 

All the benchmark doses for developmental endpoints were calculated from the Iraqi cohort data. The

adult paresthesia benchmark dose was calculated from data presented in Bakir et al. (1973). The studies

and their use in the calculation of the RfD for methylmercury are described in detail in chapters 3 and 6

of Volume IV of this Report. 

The approach used in this analysis and the EPA's RfD methodology presuppose the existence of

thresholds for certain health effects. The RfD is defined by the U.S. EPA (U.S. EPA, 1995) as 

an estimate (with uncertainty spanning perhaps an order of magnitude) of a daily

exposure to the human population (including sensitive subgroups) that is likely to be

without an appreciable risk of deleterious effects during a lifetime. 

This definition implies that the RfD is an exposure level that is below the threshold for adverse effects in

a sensitive subpopulation. For purposes of this analysis, the human population threshold is defined as the

threshold for the most sensitive individual of an identified sensitive subpopulation. The definition of

sensitive subpopulations excludes hypersensitive individuals whose susceptibilities fall far outside the

normal range. A threshold is defined as the level of exposure to an agent or substance below which a

specific effect is not expected to occur. The definition of threshold does not include concurrent exposure

to other agents eliciting the same effect by the same mechanism of action. In other words, there is an

assumption that the induced response is entirely a result of exposure to a single agent. The adverse

health endpoints for the methylmercury RfD as determined by the RfD/RfC Workgroup are the specific

clinically-observed endpoints reported in Marsh et al. (1987). The uncertainty analysis was confined to

those endpoints. The 81 pregnant female/offspring pairs comprising the Iraqi cohort were taken as a

surrogate for the most sensitive subpopulation expected in the general U.S. fish consuming population. 

The sensitive subpopulation was specifically identified as humans exposed to methylmercury in utero.

Other analyses of the Iraqi cohort data are available in the literature but are not directly

applicable to the estimation of threshold distributions. An analysis presented in the Seafood Safety

report (NAS, 1991) groups the Iraqi cohort observations by ranges of measured mercury concentrations

in hair in order to estimate the cumulative response distribution. The response data grouped by hair

mercury concentrations groupings were used to calculate the benchmark dose levels on which the

methylmercury RfD was based. As any grouping of data introduces an additional level of uncertainty,

this threshold analysis was based on the ungrouped observations. 
D-2

Cox et al. (1989) presented estimates of thresholds based on the ungrouped observations of the

Iraqi cohort for two of the five developmental endpoints considered by the U.S. EPA in the derivation of

the methylmercury RfD. Cox et al. (1989) used a threshold model that included the threshold as a

parameter. The value of the threshold parameter was estimated by a statistical method that optimized the

likelihood at different values of the threshold. The estimated threshold for late walking in offspring (first

walking after 18 months) was 7.3 ppm mercury in hair with an upper 95% confidence limit of 14 ppm. 

This threshold value was based on the best (optimized likelihood) estimate for background incidence of

late walking of 0%. The upper 95% confidence limit was highly sensitive to the value of the background

parameter, increasing to 190 ppm mercury in hair for a background of 4%. The optimized likelihood

threshold for neurological effects (neurological scores > 3) based on a background incidence of 9% was

10 ppm mercury in hair with an upper 95% confidence limit of 287 ppm. 

 

The analysis examined the major sources of uncertainty explicitly and implicitly inherent to the

methylmercury RfD and attempted to bound them quantitatively. There are a number of sources of

uncertainty in the estimation of either a human threshold or an RfD from the Iraqi cohort data and from

the dose conversion used to estimate ingestion dose levels from hair mercury concentrations. The

principal uncertainties arise from the following sources: the variability of susceptibilities within the Iraqi

cohort; population variability in the pharmacokinetic processes reflected in the dose conversion; response

classification error; and exposure classification error. 

The data show a very broad range of susceptibilities in the 81 individuals of the Iraqi cohort. An

analysis of the response rates based on hair mercury concentrations showed up to a 10,000-fold span

between the 5 and 95 percentiles when projected to the general population (Hattis and Silver, 1994). th th

Uncertainty in threshold estimates arising from the variability in individual susceptibilities was estimated

by calculating a distribution of thresholds from a regression model for repeated bootstrap samples of the

original Iraqi cohort data set. The bootstrap procedure and regression model are described in section

D.2.1. The bootstrap procedure results in a distribution of population thresholds for specific effects in

units of ppm mercury in hair. 

The methylmercury RfD used a dose conversion formula (section 6.3.1.1 of Volume IV of this

report) to estimate the ingestion dose in mg methylmercury per kg body weight per day (mg/kg-day) that

would result in a specified mercury concentration in hair. This formula comprises a number of variables

that are associated with biological processes. There are measured ranges for each variable which can be

attributed to interindividual variability in pharmacokinetics and to experimental variation. 

The response classification is the assignment of an individual observation to one of two 

categories -- responder or nonresponder. The response classification for each of the developmental

endpoints reported in Marsh et al. (1987) is based on a fixed value (response decision point) that

constitutes a response when exceeded. It is possible that some observations, particularly those that

represent responses in the immediate vicinity of the response decision point, were misclassified; a

responder may have been classified as a nonresponder or vice versa. The response classifications for late

walking and late talking are particularly susceptible to this type of error. The response estimates were

based on subject recall in members of a population that does not traditionally record these events. The

classification of neurological test battery scores is more objective but still susceptible to some degree of

investigator interpretation and misclassification.

Exposure classification error is the inclusion of individuals in the exposure group who had been

exposed outside a critical period. This type of error is a source of uncertainty for all developmental

endpoints that have a critical period of exposure combined with uncertainty about the actual timing of the
D-3

gestational period. The result of this type of error is the misclassification of an unexposed individual as

an exposed individual. The consequence of this misclassification is an overestimation of the exposure

level associated with a given response or nonresponse and subsequent overestimation of population

variability. For example, in the Iraqi cohort it is noted that an individual with the highest estimated

mercury exposure is a non-responder for developmental effects on the nervous system. This may be due

to differences in individual susceptibility to methyl mercury toxicity, or it may be a consequence of

misclassification; the individual may have been exposed during a period of time which is not critical to

development. There is potential for misclassification as the determination of the correspondence of 

gestational period and exposure was dependent on subject recall. Data pertaining to this type of

uncertainty are not yet available.

Other areas of uncertainty are those directly related to the RfD methodology. Specifically, it was

concluded by an Agency Work Group that there were no adequate chronic or reproductive studies. An

uncertainty factor of 10 is generally applied when chronic studies are not available. This uncertainty

factor is based on an assumption inherent to the RfD methodology that increased exposure duration will

lower the dose required for observation of the effect. Support for this assumption has been published

(Weil and McCollister, 1963; Dourson and Stara, 1989) and is discussed in section D.2.2.2 of this

Appendix. An uncertainty factor of 3 is generally applied if reproductive studies are not available. NoObserved-Adverse-Effect Levels (NOAELs) for reproductive studies are generally 2-fold to 3-fold higher

than NOAELs for chronic studies and are not expected to be the basis for the RfD more than 5% of the

time (Dourson, Knauf and Swartout, 1992).

D.2 Methods

Thresholds were estimated in a two-stage process. The first stage was the estimation of

threshold distributions based on hair mercury concentrations, which was accomplished by applying a

regression model to successive bootstrap samples of the observations in Marsh et al. (1987). This

process is detailed in section D.2.1. The second stage was the conversion of the thresholds expressed as

ppm mercury in hair to mg methylmercury per kg body weight per day (mg/kg-day); this involved a

Monte Carlo analysis of the variability of the underlying biological processes. The dose-conversion

model is described in section D.2.2.

For the uncertainty analysis thresholds for four of the six endpoints evaluated for the

methylmercury RfD and for combined developmental effects were estimated. The developmental effects

included in the threshold analysis were late walking, late talking and neurological effects. Thresholds for

seizures and mental symptoms were not estimated because these effects occurred at 5-fold higher hairmercury concentrations than did the other effects. As the resulting thresholds would be much higher than

the others they would not be expected to contribute significantly to the combined developmental effects

threshold distribution as defined for this analysis (the lowest of the individual effect thresholds for each

bootstrap sample). Response rates for seizures and mental symptoms would be expected to influence the

benchmark dose, however, as the benchmark dose is a function of all responses. The data used to

estimate thresholds for adult paresthesia were not the same as those used to calculate the benchmark dose

in the derivation of the methylmercury RfD. The benchmark dose was calculated from the data presented

in Bakir et al. (1973). The threshold estimates were calculated from the Iraqi cohort data so that all

thresholds would be estimated from the same group of individuals to enable a direct comparison. The

Iraqi cohort data are summarized in Table D-1. A plus (+) in Table D-1 indicates a positive response. A

positive response for neurological effects was a neurological score greater than 3 as defined in Marsh et

al. (1987). Positive responses for late walking and late talking were 18 months and 24 months (after

birth), respectively (Marsh et al., 1987).
D-4

Table D-1

Incidence of Developmental and Adult Effects as reported in Marsh, et al., 1987

max ppm neuro test late late adult max ppm neuro test late late adult

mercury in scores > 3 walking talking pares- mercury in scores > 3 walking talking pareshair thesia hair thesia

a b

1 23 +

1 26

1 38 + +

1 45 +

1 48

1 + 52

1 + 59 + +

1 + 60 + +

1 62

1 72

1 74 + +

1 75 +

1 78 + + +

1 + 86 + +

1 98

1 + 104 +

2 114 +

2 118

2 154 + + +

2 196

2 202

2 242

2 263 +

2 269

2 294 + + +

3 336 + + +

3 339 + + + +

5 357 + + +

6 362 + +

6 376 + + +

7 399

8 404 + + +

9 405 + + +

10 418 + + +

10 443 + + +

12 468 +

12 557 + +

14 + + 568 + + +

16 + 598 + +

18 + + 674 + + +

19

 defined as first walking after 18 months a

 defined as first talking after 24 months b

All threshold calculations and Monte Carlo simulations were performed in S-PLUS (ver 3.2) for ®

Microsoft Windows (ver 3.1) on several microprocessors based on the Intel 486DX2/66 ® ® ®

microprocessors. Sensitivity analyses were performed in Crystal Ball (ver 3.0) and Excel (ver 4.0) for ® ®

Windows . ®
D-5

D.2.1 Estimation of Thresholds Based on ppm Mercury in Hair

Hair mercury concentrations at the thresholds for adult paresthesia, three developmental

endpoints and combined developmental effects were estimated from the Iraqi cohort data. Threshold

estimation was accomplished by applying a regression model to successive bootstrap samples of the 81

observations. The bootstrap method (Efron and Tibshirani, 1991; 1993) is a nonparametric approach that

can be used for estimation of confidence intervals on a given variable without assuming a specific

parametric distribution for that variable. The bootstrap method is based on the assumption that the

observed sample is a random sample of a larger population and provides an estimate of sample-size

uncertainty. The bootstrap process consists of taking a random sample of the same size as the observed

sample distribution from the original sample distribution. The sampling is conducted with replacement

of selected observations prior to the next random selection such that individual observations may appear

more than once in any given sample. In this case, the bootstrap approach was applied in order to allow

estimation of confidence intervals on the dose associated with a given response. For this analysis, 5000

bootstrap samples were generated. Thresholds, in units of ppm mercury in hair, for each endpoint were

calculated for each bootstrap sample. The output was a distribution of 5000 thresholds for each endpoint

representing the variability in the population threshold as estimated from the Iraqi cohort data. The

threshold distribution for combined effects was defined as the minimum of the single-effect threshold

hair-mercury concentrations calculated at each bootstrap iteration; this definition is based on the

assumption that the endpoints were independent. The stability of the bootstrap was evaluated by

determining the change in the 5 and 95 percentiles, and their ratio, for each doubling of the number of th th

iterations. The bootstrap was considered to be stable if successive estimates were within 5% of each

other.

The threshold model used in this analysis treated background incidence and response related to

exposure (induced response) independently. The procedure is illustrated in Figure D-1, which shows the

Marsh et al., 1987 data and regression lines for developmental neurological effects (neurological score >

3). Figure D-1 is an example of a threshold determination from a single iteration of the bootstrap

procedure. All of the response data were binary; that is, individuals were either responders or

nonresponders for a given effect. The binary responses associated with each hair mercury concentration

are indicated by a "+" at the top and bottom of the chart for responders and nonresponders, respectively. 

The mean background and fitted regression lines for induced response are shown. The threshold was

defined in the regression model as the concentration of mercury in hair corresponding to the fitted

response equal to the background incidence. This is equivalent to the point of intersection of the

background and induced response line indicated as point A in Figure D-1. This model was chosen so that

the threshold estimate was a consequence of, rather than a contribution to, the estimation of background

and induced response. Other threshold models that could have been used, such as the one used by Cox et

al. (1989), include the threshold as a parameter to be simultaneously estimated with all other model

parameters. Also the maximum likelihood approach for parameter estimation was not used because of

the apparent extreme sensitivity of the upper 95% confidence limit on the threshold to variation in the

background estimate (Cox et al., 1989). 
+

+

+ + ++ + +++ + +

+

++ + +

+ +

++ +

+

+ +++

++

+

+

++ +++ +

+

++ +

+

+

+

++

++++

+

++

+

ppm mercury in hair

p

1 5 10 50 100 500

0.0 0.2 0.4 0.6 0.8 1.0

observations

background incidence

background = 0

log-linear fit

log-probit fit

+ + + +

B

A

D-6

lognormal (GM = 250, GSD = 1.35)

pd = probability density

Figure D-1

Regression Model for Determination of Bootstrap Thresholds

In the regression model, response was regressed on the logarithm of dose using the probit

function (log-probit model). In those cases where the log-probit model-predicted responses were always

greater than background, a log-linear regression model was used to determine the threshold (point B in

Figure D-1). The log-probit and log-linear fitted regression lines are shown for one bootstrap sample in

Figure D-1. 

Hair mercury concentrations of 1 ppm were assumed to represent background exposure levels

(Katz and Katz, 1992). All other observations, the first of which was at 14 ppm for any effect, were

included in the estimation of background and induced response rate as follows. Background incidence

for each effect was estimated directly from each bootstrap sample by performing repeated regressions of

response on hair mercury concentrations, starting with the assumed background range and successively

adding data points at the next higher hair mercury concentration until the regression slope was near zero

and was least statistically significant. Background incidence was defined as the mean of the fitted values

of the resulting regression. The induced response regression slope was calculated in a similar fashion,

starting with all observations above concentrations of 12 ppm hair-mercury and successively adding data

points at the next lower hair mercury concentration until the regression slope was maximized and

statistically significant (p < 0.05). 
d  CxbxV

A x f x bw

C  Hgh

hb

b  loge 2

t1⁄2

D-7

D.2.2 Estimation of Ingestion Dose Levels in mg/kg-day

D.2.2.1 Estimation of Dose Conversion Uncertainty

 The uncertainty arising from the calculation of ingestion dose levels, in mg/kg-day, corresponding to

measured concentrations of mercury in hair was estimated through analysis of the dose conversion

formula. The formula, which estimates ingested dose levels corresponding to the measured

methylmercury concentration in hair, incorporates the formula used in the derivation of the RfD with the

inclusion of an additional term to account for the hair to blood concentration ratio for methylmercury and

the conversion of elimination constants to their equivalent half-lives. The latter was done as a matter of

convenience as most of the studies reported half-lives rather than elimination constants. The formula

used in the derivation of the RfD is described in Chapter 6 (section 6.3.1.1) of Volume IV of this report

and is reproduced here as equation 1.

(1)

where

d is the daily dietary intake in mg/kg-day,

C is the concentration of methylmercury in the blood in μg/liter,

b is the elimination constant (of methylmercury from the blood) in days , -1

V is the volume of blood in the body in liters,

A is the fraction of mercury in the diet that is absorbed,

f is the fraction of absorbed mercury that is found in the blood.

bw is body weight in kg.

Variable C in formula 1 can be related to the concentration of mercury in hair by the formula given in

equation 2.

(2)

where

Hg is the concentration of mercury in hair in ppm (μg mercury/g hair), h

hb is the hair to blood concentration ratio for methylmercury in μg mercury/g hair/(μg mercury/ml

blood).

Variable b in formula 1, which is a first-order elimination rate constant, and the clearance half-life

arerelated by the formula

given in equation 3.
d  loge 2 x Hgh x V

hb x t1⁄2 x f x A x bw

DCF  loge 2 x V

hb x t1⁄2 x f x A x bw

DCFeda  DCF

UD

D-8

(3)

where

b is the elimination constant,

log 2 is the natural logarithm of 2 (= 0.693), e

t1⁄2 is the half-life of methylmercury in the blood.

Substituting for C and b in equation 1 from equations 2 and 3, respectively, gives the formula for

ingestion levels based on mercury concentrations in hair (equation 4).

(4)

Dividing both sides of equation 4 by Hgh gives the dose conversion factor, which when multiplied by

a hair mercury concentration gives the corresponding ingestion level in mg/kg-day (equation 5).

(5)

where

DCF is the dose conversion factor in ppm mercury in hair/(mg/kg-day).

Lower levels of exposure to methylmercury were expected to be associated with the observation of

effects in adults for exposure durations longer than those observed for the Iraqi cohort (U.S. EPA, 1995;

Barnes and Dourson, 1988). The potential effect of exposure duration on the dose eliciting chronic

effects is given in equation 6.

(6)

where

DCFeda is the exposure-duration adjusted dose conversion factor in ppm mercury in

hair/(mg/kg-day),

DCF is the dose conversion factor (from equation 5),

UD is a unitless adjustment for uncertainty arising from limited exposure duration.

Monte Carlo simulations were conducted for equations 5 and 6. The output of these simulations

were used to calculate a family of ingestion threshold distributions (in mg/kg-day) for each endpoint. 

This was done by multiplying the bootstrap threshold distribution for a given endpoint by specific

percentiles of the appropriate dose conversion distribution. Each member of a family of distributions was

associated with a specific probability dependent on the relative likelihood of the DCF.
D-9

D.2.2.2 Input Variable Distributions 

Distributions were assigned to each variable in equations 5 and 6 based on the data available in

the literature. The general form of the distribution, whether triangular, normal or lognormal, was

determined by examination of the shape of the distribution of empirical data and by consideration of the

underlying biological and physical processes. A triangular distribution was used when a judgement was

made that the value of the variable fell within identifiable absolute limits. Many of the variables reflect

underlying exponential processes and would be distributed as the logarithm of the nominal values. Such

variables were described as being distributed in log space. In these cases a lognormal, or log-triangular

distribution was chosen to represent the variable. 

Determination of the distribution parameters focused on identifying the median (50 percentile) th

and extreme percentiles from the available data. The focus was on the median, rather than the mean, in

order to specify percentiles of the distribution. In general, the mean value of the Monte Carlo output is

more closely related to the median, rather than the mean, of the inputs. The extreme percentiles were

those corresponding to the lowest and highest observations and were determined by multiplying the rank

order of the observation by 100/(n + 1), where n is the total number of observations. For example, the

lowest and highest observations in a sample size of 9 define the 10 and 90 percentiles (80% confidence th th

interval), respectively. Calculating the percentiles on the basis of n + 1, rather than n, allowed for the

possibility of obtaining more extreme values in additional samples. The median and extreme percentiles

were preserved in the final distribution whenever possible by adjusting the distribution parameters

accordingly. The distribution assigned to each variable is given in Table D-2.

Table D-2

Dose Conversion Monte Carlo Simulation Input Variable Distributions

Variable Form Nominal Parameters a

Value

b

hb c lognormal 250 GM =250, GSD = 1.5 f

t1⁄2 log triangular 53 days min = 1.455, mode = 1.676, max = 2.085 (log ) 10

V d triangular 5 liters min = 3.63, mode = 5.0, max = 6.37

f e log triangular 0.059 min = -1.41, mode = -1.30, max = -0.934 (log ) g

10

A triangular 0.95 min = 0.90, mode = 0.95, max = 1.0 g

bw lognormal 55 kg GM = 55, GSD = 1.13

 median or geometric median a

 Key: GM - geometric mean, GSD - geometric standard deviation; min - absolute minimum, mode - b

most likely value, max = absolute maximum

 correlated with t [correlation coefficient (r) = -0.5] c 1⁄2

 correlated with bw (r = -0.47) d

 correlated with bw (r = +0.57) e

 μg Hg/g hair/mg Hg/l blood f

 unitless ratio g

Hair to Blood Concentration Ratio for methylmercury (hb)
hair:blood ratio

pd

100 200 300 400 500

0.0 0.001 0.003 0.005

D-10

lognormal (GM = 250, GSD = 1.35)

pd = probability density

This variable represents variation in a population of the ratio of the concentration of

methylmercury in hair to the concentration of methylmercury in blood. The distribution for this variable

was based on the EPA RfD/RfC Work Group's analysis of the available data, which is presented in

Chapter 6 (section 6.3.1.1) of Volume IV of this report. The EPA RfD/RfC Work Group has judged that

the most appropriate value for this variable lies between 200 and 300 based on results published by

Phelps et al. (1980), with 250 selected as the point estimate. The value of 250 was used as the geometric

mean of the distribution for hb. The data given in Phelps et al. (1980) were not detailed enough to allow

determination of the shape or variance of the distribution. The lognormal form for the distribution was

chosen as most representative of the empirical data reported in Sherlock et al. (1982). The geometric

standard deviation (GSD) of 1.5 was estimated in this analysis from the same data (Sherlock et al., 1982). 

The distribution is shown in Figure D-2.

A correlation coefficient (r) of -0.5 was assumed between hb and t1⁄2 in the Monte Carlo

simulation of equation 5. The amount of mercury in the hair should be at least partially dependent on

how quickly methylmercury is eliminated from the blood; that is, the faster that methylmercury is

eliminated from the blood, the greater would be the difference between the concentration of mercury in

the hair and mercury in the blood. The relationship between t1⁄2 and hb would be expected to be inverse

(negative correlation); high values of t1⁄2 would be associated with low values of hb. The magnitude of

the correlation coefficient was judged by the U.S. EPA to be at least as strong as -0.5. The data available

for the calculation of the correlation between t1⁄2 and hb are extremely limited. A correlation coefficient

between hair:blood concentration and half-life of about -0.3 was calculated in this analysis from data on

four individuals (Kershaw et al., 1980). 

Figure D-2

Probability Density Function for Hair-to-Blood Concentration Ratio (hb)
D-11

Half-Life of Methylmercury in the Blood (t1⁄2)

Several human studies reported clearance half-lives for methylmercury from blood in the range

of 32-105 days with averages of 45-70 days (Miettinen et al., 1971; Bakir et al., 1973; Greenwood et al.,

1978; Kershaw et al., 1980; Smith et al., 1994). An average elimination constant for methylmercury

from the blood of 0.014 with a range of 0.0099 to 0.0165 was reported by Sherlock, et al. (1984)

corresponding to an average half-life of 50 days with a range of 42-70 days. Table D-3 gives the average

and range of reported half-lives or half-lives calculated from equation 3 for each of the five studies. The

values in Table D-3 were for male and female adults, combined. Average half-lives of methylmercury in

the blood, as reported by Greenwood, et al. (1984), were somewhat shorter for lactating women (46 days)

and longer for children (90 days) than for the adult average of 70 days. A histogram of the combined

data was highly skewed and roughly triangular in shape. The log-triangular distribution (Figure D-3) was

chosen as best representative of the empirical data. The median value of this distribution was 53 days;

this was slightly higher than that used in the derivation of the methylmercury RfD (50 days), which

would result in slightly lower dose conversion values.

Table D-3

Half-Life of Methylmercury in the Blood (days)

Reference Low Average High

Smith et al. 1994 31.9 45.3 60

Sherlock et al. 1984 42 49.5 70

Kershaw et al. 1980 46.7 51.9 66.5

Bakir et al. 1973 40 65 105

Greenwood et al. 1978 49 70 95
half-life (days)

pd

30 40 50 60 70 80 90 100

0.0 0.5 1.0 1.5

D-12

log triangular {min = 1.455, mode = 1.676, max = 2.085 (log10)}

pd = probability density

Figure D-3

Probability Density Function for the Half Life of Methylmercury in the Blood (t 1⁄2)

Volume of Blood in the Body (V)

This variable represents the variation in a population of the total volume of blood in the body. 

The distribution was based on published values of estimated whole blood volumes for a cohort of 20

pregnant Nigerian women (Harrison, 1966). Whole blood volumes in the third trimester of pregnancy

ranged from 4.0 to 6.0 liters ; the mean and median values were both 5.0 liters (Harrison, 1966). The

distribution of empirical data was roughly triangular and symmetrical. The minimum and maximum

values were adjusted so that the range of observed values fell within the 90% confidence interval (n =

20). The distribution is shown in Figure D-4.

Blood volume was assumed to be positively correlated with body weight; larger blood

volumes would be associated with higher body weights. For this analysis a correlation coefficient of

0.57 between V and bw from the data given in Harrison (1966) was calculated. This correlation

coefficient was assumed for the standard Monte Carlo simulation of equation 5. 
blood volume

pd

0.035 0.040 0.045 0.050 0.055 0.060 0.065

0 20 40 60 80

D-13

triangular {min = 3.5 liters, mode = 5.0 liters, max = 6.5 liters}

pd = probability density

Figure D-4

Probability Density for Blood Volume (V)

Fraction of Absorbed methylmercury in the Blood (f)

This variable reflects the distribution and dilution of the absorbed methylmercury in all

compartments of the body. The distribution on f was based on several human studies showing values in

the range of 5-10% of the absorbed dose of methylmercury in the blood (Miettinen et al., 1971; Kershaw

et al., 1980; Sherlock et al., 1984). All of the measured values have been adjusted for an assumed total

blood volume of 5 liters. The studies are summarized in Chapter 6 (section 6.3.1.1) of Volume IV of this

report. The distribution is shown in Figure D-5. The median value of this distribution of 5.9% was

higher than that used in the derivation of the methylmercury RfD (5%), which would result in lower dose

conversion values.

Sherlock et al. (1984) reported that the fraction of methylmercury in the blood was negatively

correlated with body weight; smaller fractions of methylmercury in the blood were associated with

larger body weights. For this analysis the U.S. EPA calculated a correlation of -0.47 between f and bw

from the data given in Sherlock et al. (1984). This correlation was assumed for the Monte Carlo

simulation of equation 5. 
fraction MeHg in the blood

pd

0.04 0.05 0.06 0.07 0.08 0.10

0.0 0.5 1.0 1.5 2.0

absorption factor

pd

0.90 0.92 0.94 0.96 0.98 1.00

0 5 10 15 20

D-14

log triangular {min = -1.41, mode = -1.30, max = -0.934 (log10)}

pd = probability density

triangular {min = 0.90, mode = 0.95, max = 1.0}

pd = probability density

Figure D-5

Probability Density for Fraction of Absorbed Methylmercury in the Blood (f)

Fraction of methylmercury in the Diet that is Absorbed (A)

This distribution was based on the results of two human studies showing uptake of radiolabeled methylmercury of 95% and greater (Aberg et al., 1969) and 94% and greater (Miettinen et al.,

1971) and animal studies showing 90% or greater absorption (summarized in Walsh, 1982). The

distribution reflects the expectation that this value is close to 100 and will not vary much. The

distribution is shown in Figure D-6.

Figure D-6

Proba

bility

Densit

y for

Fracti

on of

Methy

lmerc

ury

Absor

bed

from

the

Gut

(A)
body weight

p dens

40 50 60 70 80

0.0 0.01 0.02 0.03 0.04 0.05 0.06

D-15

lognormal (GM = 55 kg, GSD = 1.13)

pd = probability density

Body Weight (bw)

The distribution for body weight was based on Harrison (1966), previously described for the

definition of V. The observed body weights during the third trimester of pregnancy ranged from 49.5 kg

to 73.9 kg with a geometric mean of 55 kg (Harrison, 1966). A lognormal distribution was visually

fitted to the data. The distribution is shown in Figure D-7. The median value of this distribution of 55 kg

was lower than that used in the derivation of the methylmercury RfD (60 kg). Use of the lower value for

bw which would result in higher dose conversion values. 

Figure D-7

Probability Density Distribution for Body Weight (bw)

Uncertainty Arising from Limited Exposure Duration (UD)

This factor is an adjustment for the uncertain effects of exposure duration on the magnitude of

the effective dose. It is based on the assumption that continuing exposure will result in the appearance of

effects at exposure levels wherein there were no effects observed following shorter exposure durations. 

The U.S. EPA commonly applies an uncertainty factor of 10 when calculating a chronic RfD from a

study of subchronic duration (U.S. EPA, 1995). In concept, the value of 10 for this uncertainty factor

represents a high estimate of the uncertainty in order to maintain the protective nature of the RfD. An

empirical analysis of the Weil and McCollister (1963) data by Dourson and Stara (1983) supports the use

of an uncertainty factor of 10 as protective. About 50% of ratios of subchronic NOAELs to chronic

NOAELS for rats exposed to a variety of substances other than methylmercury (as reported by Weil and

McCollister, 1963) were below 3.5 and 95% were below 10 (Dourson and Stara, 1983). 

The published data were insufficient for the estimation of a distribution for UD. A point

estimate of 4.7 was made for UD from a few studies of methylmercury toxicity in nonhuman primates. 

These studies are summarized in Table D-4. Table D-4 gives NOAELs and Lowest-Observed-AdverseEffect Levels (LOAELs) for studies of short-term and long-term duration in monkeys. The neurologic

endpoints were limited to clinically-observable effects in order to maintain approximate equivalence of

effects across exposure durations. UD was estimated by dividing the short-term LOAEL of 0.21 mg/kg-
D-16

day (Sato and Ikuta, 1975) by 0.045, the average of the two long-term exposure LOAELS of 0.04 and

0.05 mg/kg-day (Burbacher et al., 1988; Rice and Gilber, 1992). 

UD was used in equation 6 to adjust the dose conversion factor for the estimation of the

exposure level associated with chronic effects. Specifically, the exposure duration adjusted dose

conversion factor (DCFeda) from equation 6 was multiplied by the adult paresthesia bootstrap threshold

distribution to obtain an ingestion threshold distribution for chronic neurologic effects. The precise

nature of the chronic effects was not specified because the effects observed in the monkey studies used to

define UD included a number of different neurologic effects. In this case the paresthesia observed in the

Iraqi cohort was used as a surrogate for all possible adult neurologic effects that might occur following

short-term exposure to methylmercury.

Table D-4

Methylmercury Toxicity in Animals

Reference Exposure Effects NOAEL LOAEL

Duration

Sato & Ikuta 1975 36-132 days ataxic gait, myoclonic 0.07 0.21

seizures

Burbacher et al., 3 years slight tremor, motor none 0.04

1988 incoordination,

blindness

Rice & Gilbert, 6.5-7 years decreased fine motor none 0.05

1992 performance, other

D.2.2.3 Correlation of Input Variables

Apart from the assumptions of correlation between individual input variables described

previously (standard correlations), a simplifying assumption was that the susceptibility of any individual

was independent of the value of the dose-conversion factor. It is very likely, however, that susceptibility

and t1⁄2 are correlated. Longer residence times of methylmercury in the blood, corresponding to longer

half-lives, should have a direct effect on toxicity. Thus, there would be some likelihood that the

susceptibility of the individual at the population threshold (the most sensitive individual) would be

related to larger values of t1⁄2 1⁄2 . Monte Carlo analyses of equation 5 limiting t to values greater than 53

days (the median of the t1⁄2 distribution) or greater than 84 days (the 90 percentile of the t1⁄2 distribution) th

were also conducted. The latter simulation was included only to determine the sensitivity of the output to

changes in the assumption and was not considered to be a realistic scenario. Standard input variable

correlations were assumed for these simulations. Results of this simulation are presented in section D.3.

A sensitivity analysis was conducted to examine the effect of different correlation assumptions

on the relative contribution of each input variable distribution to the variance of the Monte Carlo

simulation output. The sensitivity analysis was performed for standard correlations, no correlations and

for the alternate half-life (t1⁄2 1⁄2 > 53 days or t > 84 days) scenarios. 
D-17

D.2.3 Estimation of Uncertainty Arising from Response Classification Error

A variable-response model was constructed to assess response classification error. The

variable-response model is identical to the general threshold model except that responses presumed to be

uncertain were given fractional values rather than 0 or 1 (for nonresponders and responders,

respectively). Values of 0 or 1 were generated at each bootstrap sampling by comparison of the

fractional value with random numbers drawn from a uniform distribution between 0 and 1. The uncertain

observations were defined as those that fell close to the defined minimum response. For late walking the

observations that fell into this category were those of 18 - 20 months (late walking = not walking by 18

months). For late talking the uncertain observations were 24 - 26 months (late talking = not talking by 24

months). A value of 0.5 was assigned to each of the observations for late walking and late talking that

were designated uncertain; this represented the largest possible uncertainty in classification (50%

classification error). A 50% classification error was judged to be plausible, given the highly variable

factors involved in the original classifications in Marsh et al. (1987). A separate analysis was conducted

for late walking assuming a 25% classification error. This was done to allow for the possibility that the

large number of observations at exactly 18 months (22), was a result of 18 months being used as an upper

bound, rather than an exact estimate. This could have occurred for observations that were uncertain but

judged by the authors (Marsh et al., 1987) to be 18 months or less.

The determination of neurological scores in Marsh et al. (1987) was considered to be more

objective than the determination of late walking or late talking. There is, however, a possibility that the

distinction between adjacent scores is not absolute. The uncertain observations for neurological effects

were scores of 3 or 4. Although there was no clear basis for determining classification error for this

endpoint, the error was judged likely to be considerably less than for late walking and late talking. 

Simulations were run assuming a 10% or 20% classification error. A classification error rate of 20% was

considered to be an upper bound.

There was no basis on which to determine the extent of classification error for adult

paresthesia. In addition, there was no way to determine which responses (or nonresponses) were

marginal. A 5% classification error was assumed for all observations to determine the sensitivity of the

threshold simulation to small error rates. 

D.3 Results

D.3.1 Bootstrap Analysis

Figures D-8 to D-10 show the frequency distributions of thresholds for the individual

developmental endpoints resulting from the bootstrap analysis. Figure D-11 is the threshold distribution

for the occurrence of any developmental effect. The figures are histograms of the frequency of

occurrence of calculated threshold mercury concentrations resulting from 5,000 iterations of the

bootstrap procedure. The threshold values on the abscissa are given in log units. These distributions 10

represent uncertainty in the estimation of population thresholds calculated from hair mercury

concentrations. The small separate peaks at about 600 ppm in Figures D-8 to D-10 represent bootstrap

samples that result in a nonsignificant (p > 0.05) log-probit regression slope. A nonsignificant slope

implies that there was no relationship between hair mercury concentrations and observed response for

that bootstrap sample. In these cases the threshold was defined as the largest hair mercury concentration

in the sample. The frequency with which nonsignificant slopes occurred was interpreted as a measure of

the reliability of the endpoint as a measure of methylmercury toxicity. This occurred in 0.4% of the

bootstrap samples for neurological effects, 0.1% of the late walking bootstrap samples and in 0.2% of the
-1 0 1 2 3

0 100 200 300 400 500

ppm Hg in hair (log10)

frequency

5000 iterations

D-18

late walking bootstrap samples. These percentages effectively limit the upper end of the bootstrap

confidence intervals that can be defined meaningfully. As an example, the upper limit on the bootstrap

confidence interval for the neurological effects threshold was 99.6%. The maximum two-tailed

symmetrical confidence interval on the median threshold for this endpoint that excluded nonsignificant

slopes was 99.2%.

At the other end of the distributions, the log-probit model fails to estimate a threshold when the

lowest log-probit response is greater than the background incidence (log-probit regression line fails to

intersect background in Figure D-1). This would occur in all bootstrap samples where a zero background

incidence was estimated or when the calculated log-probit response at 3 ppm mercury in hair was greater

than the background incidence for that sample. The threshold was calculated from the log-linear

regression model in these cases (point B in Figure D-1); this occurred in 6%, 13.5% and 100% of the

bootstrap samples for neurological effects, late talking and late walking, respectively. The average

background incidence, as estimated from the Iraqi cohort data for each bootstrap sample, was 10.7% for

neurological effects and 8.6% for late talking. Background incidence for late walking was 0% for all

samples.

Figure D-8

Bootstrap Threshold Distribution for Developmental Neurological Effects
-1 0 1 2 3

0 100 200 300 400 500

ppm Hg in hair (log10)

frequency

5000 iterations

-1 0 1 2 3

0 200 400 600

ppm Hg in hair (log10)

frequency

5000 iterations

D-19

Figure D-9

Bootstrap Threshold Distribution for Late Walking

Figure D-10

Bootstrap Threshold Distribution for Late Talking
-1 0 1 2 3

0 100 200 300 400 500 600

ppm Hg in hair (log10)

frequency

5000 iterations

-1 0 1 2 3

0 100 200 300 400 500

ppm Hg in hair (log10)

frequency

5002 iterations

D-20

Figure D-11

Bootstrap Threshold Distribution for Combined Developmental Effects

Figure D-12 shows the distribution of bootstrap thresholds for adult paresthesia . The

distribution is a result of 5000 iterations of the bootstrap procedure. Nonsignificant regression slopes

occurred in 7.5% of the samples as shown by the peak at around 2.8 (600 ppm) in Figure D-12. The

largest confidence interval for the adult paresthesia threshold that excluded nonsignificant slopes was

85%. Background incidence for adult paresthesia was 0% for all samples.

Figure D-12

Bootstrap Threshold Distribution for Adult Paresthesia
D-21

Table D-5 gives selected percentiles from the cumulative bootstrap threshold distributions for

the developmental endpoints and adult paresthesia. The threshold values given in Table D-5 are given in

units of ppm mercury in hair. The values given for the 5 and 95 percentiles in Table D-5 define the th th

90% bootstrap confidence interval for each threshold. The adult paresthesia thresholds were the lowest

of all the endpoints modeled but showed the greatest variability. The late walking threshold was the

lowest and most variable of the individual developmental endpoint thresholds. The combined-effects

threshold was the least variable of all the thresholds as would be expected from the method of calculation

(minimum of the three individual endpoint thresholds). For the combined developmental effects, the late

walking threshold was the lowest of the three thresholds most often (45%) with neurological effects and

late talking contributing the lowest threshold 31% and 24% of the time, respectively. 

Table D-5

Bootstrap Threshold Distributions in ppm Mercury in Hair for All Effects 

Bootstrap Percentile

Endpoint 5 25 50 75 95 th th th th th

neurological effects 3.8 10 19 33 63 a

late walking 3.6 8.0 14 25 58 b

late talking 5.5 13 20 31 57 c

combined developmental effects 24 2.5 5.3 8.7 14 24 4

adult paresthesia 0.64 1.5 2.8 5.9 > 500

 neurological test scores > 3 in children exposed in utero a

 walking after 18 months b

 talking after 24 months c

 threshold for the occurrence of any developmental effect d

All bootstraps stabilized within 4000 iterations as measured by the change in the 5 and 95 th th

percentiles and the ratio of those percentiles. The largest change from 2000 to 4000 iterations in any of

the stability measurements was 3.5%. 

D.3.2 Response Classification Uncertainty 

Table D-6 gives percentiles of the cumulative bootstrap threshold distributions resulting from

the consideration of response classification error for each of the endpoints. The distributions were a

result of 5000 iterations of the bootstrap procedure. Frequency plots for these distributions are shown in

Figures D-13 and D-14 for late walking and adult paresthesia, respectively.
-1 0 1 2 3

0 100 200 300 400 500

ppm Hg in hair (log10)

frequency

5000 iterations

D-22

Table D-6

Bootstrap Threshold Distributions in ppm Mercury in Hair with Inclusion of

Response Classification Error 

Bootstrap Percentile

Endpoint 5 25 50 75 95 th th th th th

neurological effects 2.6 8.4 16 30 71

(CE = 10%) a

neurological effects 2.3 7.4 15 30 > 600

(CE = 20%)

late walking 0.74 2.5 5.7 15 > 600

(CE = 25%)

late walking 0.79 2.6 5.0 12 > 600

(CE = 50%)

late talking 2.1 5.9 12 25 99 b

adult paresthesia 0.50 1.5 3.3 15 > 600 c

 classification error assumption for responses at boundary of minimum value defining a positive response a

 50% classification error (boundary responses) b

 5% classification error assumed for all responses above background c

Figure D-13

Bootstrap Threshold Distribution for Late Walking with Response-Classification Error
-1 0 1 2 3

0 100 200 300 400 500

ppm Hg in hair (log10)

frequency

5000 iterations

D-23

Figure D-14

Bootstrap Threshold Distribution for Adult Paresthesia with Response-Classification Error

The primary result of the assumptions of response classification error was an increase in the

number of bootstrap samples resulting in a nonsignificant log-probit regression slope as shown in Table

D-7. Late walking and adult paresthesia, for which over 20% of the regression slopes were

nonsignificant, were the most sensitive to classification error. Bootstrap confidence intervals of less than

60% were the largest that excluded nonsignificant log-probit slopes for both endpoints. 6.1% of the

slopes for neurological effects were nonsignificant when a 20% classification error was assumed; there

was little effect when the error estimate was reduced to 10%. Only 4.6% of the slopes for late walking

were nonsignificant with a classification error of 50%; the width of the 90% confidence interval,

however, increased by a factor of 4. Eliminating late walking from the combined developmental effects

resulted in a 53% increase in the median bootstrap threshold to 12 ppm; the 5 and 95 percentiles were th th

increased to 2.7 and 32 ppm, respectively.
D-24

Table D-7

Percentage of Bootstrap Thresholds Resulting from 

Nonsignificant Log-Probit Regression Slopes 

Endpoint standard classification error 1

regression regression

neurological effects 0.4 CE = 10% 2.1 2

CE = 20% 6.1

late walking 0.8 CE = 25% 10.9

CE = 50% 21.9

late talking 0.2 CE = 50% 4.6

adult paresthesia 7.5 CE = 5% 20.9 3

 no classification error a

 classification error assumption b

 all observations c

D.3.3 Dose Conversion Monte Carlo Simulation and Sensitivity Analysis

Table D-8 shows the results of the Monte Carlo simulation of the dose conversion factor (DCF

in equation 5) for different correlation assumptions. Standard assumptions (scenario 1, Table D-8) were

the distribution assignments and correlations described in section D.2.2.2 and summarized in Table D-2. 

Scenario 2 in Table D-8 assumed that all the variables in equation 5 were independent. Scenarios 3 and

4 included the standard assumptions and an additional assumption that an increased residence time of

methylmercury in the blood contributed to the susceptibility of the most sensitive individual. Scenario 3

restricted t1⁄2 to the upper half of the standard distribution, representing a moderate association of half-life

and susceptibility, while scenario 4 represented a stronger association, restricting t 1⁄2 to the upper 10% of

the standard distribution. 

Figure D-15 shows the dose conversion frequency distributions for the standard dose

conversion factor distribution (scenario 1, Table D-8). The values on the abscissa are given in ppm

mercury in hair/(mg/kg-day) in log units. The distributions in Table D-8 and Figure D-15 represent the 10

uncertainty in the ratio of the exposure level (in mg/kg-day) to hair mercury concentration for the most

sensitive individual of the exposed population. The nominal dose conversion factor is defined here as the

median value of the standard simulation (scenario 1, Table D-8). The median value for the standard

simulation was 8.0 x 10 with a 90% confidence interval spanning a 3.57-fold range. The corresponding -5

dose conversion value used in the derivation of the methylmercury RfD was 9.8 x 10 . That is, the -5

methylmercury RfD would change very little if calculated using the median of the simulated dose

conversion distribution. Using the nominal dose conversion factor, an exposure level of 1 x 10-4

mg/kg-day corresponds to a hair mercury concentration of 1.25 ppm, with a 90% confidence interval of

0.69 ppm to 2.36 ppm.
-5.0 -4.5 -4.0 -3.5

0 200 400 600 800 1200

mg/kg-day (log10)

frequency

10000 iterations

D-25

Monte Carlo simulation of equation 4 with HgH = 1

Standard assumptions as in Table D-2

Table D-8

Dose Conversion Factor Monte Carlo Simulation Output for Different Correlation Assumptions in

mg/kg-day 

Percentile

Scenario 5 25 50 75 95 th th th th th

1) standard correlations 4.2 x 10 6.2 x 10 8.0 x 10 1.0 x 10 1.5 x 10 a -5 -5 -5 -4 -4

2) no correlations 2.6 x 10 5.0 x 10 7.9 x 10 1.2 x 10 2.3 x 10 -5 -5 -5 -4 -4

3) t1⁄2 > 53 days (std. correlations) b 3.0 x 10 4.7 x 10 6.2 x 10 8.3 x 10 1.3 x 10 -5 -5 -5 -5 -4

4) t1⁄2 > 84 days (std. correlations) c 2.2 x 10 3.4 x 10 4.6 x 10 6.1 x 10 9.4 x 10 -5 -5 -5 -5 -5

 hb correlated with t (r = -0.5); f correlated with bw (r = -0.47); V correlated with bw (r = +0.57) a 1⁄2

 50 percentile of t b th 1⁄2

 90 percentile of t c th 1⁄2

Figure D-15

Dose Conversion Distribution (standard assumptions)

The result of assuming correlations between input variables in equation 5 (Table D-8, scenario 1)

is a 60% reduction of the width of the 90% confidence interval compared to assuming total independence

of inputs (scenario 2). Conversely, restricting t1⁄2 to the upper half of the distribution (correlating

susceptibility and t1⁄2) resulted in increased uncertainty around the DCF and lower dose conversion

estimates. Reductions in the median dose conversion estimate were 22% and 42% for a moderate (t 1⁄2 >

53 days) and a strong (t1⁄2 1⁄2 > 84 days) association of t and susceptibility, respectively; 
D-26

the increase in the width of the 90% confidence interval, as measured by the ratio of the 95 and 5 th th

percentiles, was about 20% in both cases. 

Table D-9 shows the relative contribution of each dose conversion input variable to the variance

of the Monte Carlo simulation output for selected scenarios from Table D-8. It can be seen from Table

D-9 that hb contributed the most to the variance of the output across the scenarios, while bw, V and A

contributed relatively little. The relative contribution to the output variance of t1⁄2 and f was highly

sensitive to the correlation assumptions.

Table D-9

Sensitivity Analysis for Dose Conversion Monte Carlo Simulation: 

Contribution of Each Input Variable to Output Variance (%)

Scenario

Input Variable no correlations standard

correlationsa

alternate t1⁄2

b

hb 47.9 46.5 60.4

t1⁄2 26.7 7.9 0.0

f 16.3 33.3 29.0

V 4.7 9.6 8.3

bw 4.3 2.4 1.9

A 0 0.3 0.4

 hb correlated with t (r=-0.5); f correlated with bw (r=-0.47); V correlated with bw (r=+0.57) a 1⁄2

 t > 53 days b

1⁄2

The variability of the dose conversion simulation was somewhat less than the contribution from

the bootstrap procedure. The widths of the 90% bootstrap confidence intervals on the thresholds (in ppm

mercury in hair) ranged from 1.1 to 1.3 orders of magnitude (12-20 fold difference in the 5 and 95 th th

percentiles from Table D-5). The width of the 90% confidence interval for the standard dose conversion

simulation spanned 0.55 orders of magnitude (Table D-8), or about 18-30% of that for the bootstrap

confidence intervals. 

D.3.4 Estimation of Ingestion Thresholds

The distributions given in Table D-7 were used to obtain dose-conversion confidence intervals

for specific threshold estimates. Table D-10 gives values at selected percentiles for the distribution of

dose-conversion uncertainty around the median ingestion threshold estimates. The values in Table D-10

are given in units of 10 mg/kg-day as a convenience for comparison with the RfD of 1 x 10 mg/kg-day. -4 -4

The distributions in Table D-10 were determined by multiplying the appropriate dose conversion

distribution from Table D-7, as noted in Table D-10, by the median bootstrap threshold estimates for

each of the endpoints given in Table D-5. That is, the distributions in Table D-10 represent the output of

equations 5 or 6 with Hgh equal to the median of the indicated bootstrap distribution. As an example, the

distribution for developmental neurological effects in Table D-10 was a result of multiplying the 5 , 25 , th th

50 , 75 and 95 percentiles of the standard DCF distribution by the median bootstrap threshold (19 th th th
D-27

ppm mercury in hair) for developmental neurological effects. The duration-adjusted adult paresthesia

distribution was the dose conversion distribution for adult paresthesia given in Table D-11 divided by 4.7

(UD). Table D-11 is the equivalent of Table D-8 for the 5 percentile bootstrap threshold estimates from th

Table D-5. 

Table D-10

Dose Conversion Distributions for Median Ingestion Threshold Estimates 

in mg/kg-day (x 10 )4

Percentile

Endpoint 5 25 50 75 95 th th th th th

neurological effects 7.9 12 15 19 27 a

late walking 6.1 9.0 12 15 21 a

late talking 8.4 12 16 20 29 a

combined developmental effects 3.7 5.4 7.0 8.7 13 a

combined developmental effects (t 2.6 4.0 5.4 7.2 11 1⁄2

> 53 days)b

adult paresthesia 1.2 1.7 2.2 2.8 4.0 a

duration-adjusted adult paresthesia 0.26 0.36 0.47 0.60 0.86 c

 standard assumptions (scenario 1, Table D-9) a

 scenario 3, Table D-9 b

 adult paresthesia distribution divided by U c D

Table D-11

Dose Conversion Distributions for 5 Percentile Ingestion Threshold Estimates th

in mg/kg-day (x 10 )4

Percentile

Endpoint 5 25 50 75 95 th th th th th

neurological effects 1.6 2.4 3.1 3.8 5.6 a

late walking 1.5 2.2 2.9 3.6 5.2 a

late talking 2.3 3.4 4.4 5.5 8.0 a

combined developmental effects 1.0 1.5 2.0 2.5 3.6 a

combined developmental effects 0.74 1.1 1.5 2.0 3.1 b

(t > 53 da 1⁄2 ys)

adult paresthesia 0.27 0.40 0.52 0.65 0.94 a

duration-adjusted adult paresthesia 0.058 0.086 0.11 0.14 0.20 c

 standard assumptions (scenario 1, Table D-9) a
D-28

 scenario 3, Table D-9 b

 adult paresthesia distribution divided by U c D

The 5% and 95% columns in Tables D-10 and D-11 represent the 90% confidence intervals for

specific percentiles of the thresholds expressed as ingestion levels in mg/kg-day. For example, with

respect to the neurological effects distribution in Table D-10, there is 90% confidence that the true

median threshold for neurological effects is between 7.8 x 10 and 2.7 x 10 mg/kg-day. Similarly, there -4 -3

is 90% confidence that the true 5 percentile of the neurological effects threshold distribution is between th

1.4 x 10 and 4.8 x 10 mg/kg-day (Table D-11). The median ingestion threshold for duration-adjusted -4 -4

adult paresthesia was 1 x 10 mg/kg-day, with a 90% confidence interval of 2.3 x 10 mg/kg-day to 4 x -4 -5

10 mg/kg-day. -5

Figure D-16 is a plot of the cumulative bootstrap threshold distribution for combined

developmental effects multiplied by values of the dose conversion distribution at selected percentiles. 

The plots from left to right in Figure D-16 represent different realizations of the distribution of ingestion

thresholds based on the relative likelihood of specific values of the dose conversion factor (5 , 50 and th th

95 percentiles). The horizontal box and whisker plot corresponds to the dose conversion distribution th 

multiplied by the median of the bootstrap threshold distribution for combined developmental effects as

given in Table D-10; the box is the interquartile range (25 to 75 percentiles) and the whiskers are the th th

5 and 95 percentiles. Figure D-17 is the same plot for combined developmental effects with the th th

assumption that t1⁄2 is greater than 53 days (Table D-8, scenario 3). Figures D-18 to D-22 are the

equivalent plots for the individual-effect thresholds. Ingestion threshold distributions for adult effects

are shown in Figures D-21 and D-22. Figure D-21 is the ingestion threshold distribution for the adult

paresthesia observed in the Iraqi cohort. Figure D-22 is the ingestion threshold distribution for durationadjusted adult paresthesia resulting from dividing the adult paresthesia ingestion threshold distribution by

UD. That is, Figure D-22 is the distribution in Figure D-21 shifted to the left by a factor of 4.7.

Figure D-16

Cumulative Combined Developmental Effects Ingestion Threshold Distribution
threshold (mg/kg-day)

p

0.00005 0.00050 0.00500

0.0 0.2 0.4 0.6 0.8 1.0

0.0001 0.001

ingestion threshold distribution

dose conversion distribution

RfD

5% 25% 50% 75% 95%

RfD

threshold (mg/kg-day)

p

0.00005 0.00050 0.00500

0.0 0.2 0.4 0.6 0.8 1.0

0.0001 0.001

ingestion threshold distribution

dose conversion distribution

RfD

5% 25% 50% 75% 95%

RfD

D-29

p = cumulative probability

p = cumulative probability

Figure D-17

Cumulative Combined Developmental Effects Ingestion Threshold Distribution

(t > 53 days 1⁄2 )

Figure D-18

Cumulative Neurological Effects Ingestion Threshold Distribution
threshold (mg/kg-day)

p

0.00005 0.00050 0.00500

0.0 0.2 0.4 0.6 0.8 1.0

0.0001 0.001

ingestion threshold distribution

dose conversion distribution

RfD

5% 25% 50% 75% 95%

RfD

threshold (mg/kg-day)

p

0.00005 0.00050 0.00500

0.0 0.2 0.4 0.6 0.8 1.0

0.0001 0.001

ingestion threshold distribution

dose conversion distribution

RfD

5% 25% 50% 75% 95%

RfD

D-30

p = cumulative probability

p = cumulative probability

Figure D-19

Cumulative Late Walking Ingestion Threshold Distribution

Figure D-20

Cumulative Late Talking Ingestion Threshold Distribution
threshold (mg/kg-day)

p

0.00005 0.00050 0.00500

0.0 0.2 0.4 0.6 0.8 1.0

0.0001 0.001

ingestion threshold distribution

dose conversion distribution

RfD

5% 25% 50% 75% 95%

RfD

threshold (mg/kg-day)

p

0.00001 0.00005 0.00050

0.0 0.2 0.4 0.6 0.8 1.0

0.0001 0.001

ingestion threshold distribution

dose conversion distribution

RfD

5% 25% 50% 75% 95%

RfD

D-31

p = cumulative probability

p = cumulative probability

Figure D-21

Cumulative Adult Paresthesia Ingestion Threshold Distribution

(no exposure duration adjustment)

Figure D-22
threshold (mg/kg-day)

p

1e-6 5e-6 5e-5 5e-4

0.0 0.2 0.4 0.6 0.8 1.0

1e-4 1e-3

ingestion threshold distribution

dose conversion distribution

RfD

5% 25%50%75% 95%

RfD

D-32

p = cumulative probability

Cumulative Duration-Adjusted Adult Paresthesia Threshold Distribution

The uncertainty around the location of the RfD within each of the distributions shown in Figures

D-17 to D-22 is indicated by the vertical line at 1 x 10 mg/kg-day; this uncertainty came from the dose -4

conversion variability. As an example, the RfD fell between the 0.035 and 4.5 percentiles of the

combined developmental effects threshold distribution with 90% confidence as determined by the

intersection of the RfD line with the 5 % and 95% ingestion threshold curves (Figure D-16). The median

estimate of the location of the RfD in this distribution was the 0.25 percentile. Similarly, the RfD fell

between the 39 and 91 percentile of the duration-adjusted adult paresthesia threshold distribution with th st

the median at the 75 percentile (Figure D-22). The RfD was at the 18 percentile of the adult th th

paresthesia threshold distribution and below the 1 percentile for all of the other threshold distributions. st

D.4 Discussion of Uncertainty Analysis

Because the Iraqi cohort is considered to be a sensitive subgroup, as defined in the RfD

methodology, the output distributions of the analysis are meant to reflect the uncertainty around an

estimate of the thresholds for effects in humans including sensitive individuals. The results for each

endpoint should be interpreted as the distribution of the uncertainty around the human population

threshold. The results should not be interpreted as the distributions of individual thresholds within the

population. Estimates of risk above the threshold cannot be obtained from this analysis.

This analysis has attempted to incorporate all areas of uncertainty involved in the derivation of

the methylmercury RfD in Chapter 6 of Volume IV of this report. The 10-fold uncertainty factor (UF)

includes a 3-fold (10 ) factor for human variability and a 3-fold factor for the lack of reproductive and 0.5

chronic studies. The UF for human variability includes a consideration of both susceptibility and

variation in methylmercury blood half-lives. The bootstrap threshold confidence intervals incorporate

the former, and the latter is explicitly modeled (t1⁄2) in the dose conversion Monte Carlo analysis. 

Uncertainty arising from the lack of chronic data is estimated by UD; this uncertainty was a point estimate

only, as the data were inadequate for defining a distribution for UD D . Because U was derived as a scaling
D-33

factor for adult effects it is not directly comparable to the UF for chronic effects used in the derivation of

the RfD, which was based on developmental endpoints. The only uncertainty included in the RfD and

not addressed here is the uncertainty attributed to the lack of a reproductive study; there are no

appropriate data for the estimation of this uncertainty. In general, reproductive NOAELs are slightly

lower than developmental NOAELs for other substances, but much higher than chronic NOAELs

(Dourson, Knauf and Swartout, 1992). That is, the uncertainty in the thresholds is expected to be much

less for lack of a reproductive study than for lack of a chronic study.

The uncertainty analysis presented in this appendix was limited to only those data and formulae

directly related to the derivation of the methylmercury RfD. Other data sets or models were not

considered. A few sources of uncertainty in the data used to derive the methylmercury RfD have not

been included in this analysis. Exposure classification error arising from uncertainty as to the

correspondence of actual exposure and critical exposure period cannot be estimated from the data as

published in Marsh et al., 1987. This source of uncertainty could be a major contributor to the apparent

extreme variability of susceptibilities in the Iraqi cohort. Variability in the interpretation of the definition

of a response was not estimated in this analysis. That is, there would have been some differences in how

individuals interpreted what constituted first walking or first talking, probably more so for the latter. The

classification errors assumed for this analysis only account for uncertainty in the timing of the event

given an unequivocal positive response. Also, the response decision points defining an adverse effect

were accepted uncritically. For example, changing the definition of late walking to either greater than 16

months or greater than 20 months would have a significant effect on the analysis. Measurement error for

hair mercury concentrations has not been estimated for this analysis; the necessary data are unavailable

in the published reports (Marsh et al., 1987; Cox et al., 1989). In addition, the results of this analysis are

conditional on a specific representation of population variability in the parameters of the dose conversion

variables. That is, the choice of the form, and parameters for the distributions assigned to each of the

variables is largely a matter of judgement. The particular set of parameters chosen for each distribution

is only one option of a number of possible choices; uncertainty as to the value of the parameters is not

included in the analysis. 

The threshold analysis shows that adult paresthesia was the most sensitive individual effect

observed for the Iraqi cohort, particularly when adjusted for the effects of continuing exposure. That is,

in this analysis, paresthesia in adults was estimated to be observable at a lower exposure than the

developmental endpoints. The absence of an observed background incidence for paresthesia in the Iraqi

cohort partially contributed to the low threshold estimates. A background incidence for paresthesia

would be expected in the general population. The adult paresthesia bootstrap thresholds were also the

most unstable as measured by the frequency of nonsignificant slopes. The RfD fell between the 39 and th

91 percentiles of the duration-adjusted adult paresthesia threshold distribution, a considerably larger st

range than that for any of the developmental effects. On the average, the RfD fell below the 1 percentile st

for all developmental effects, with only a 5% chance that it was as high as the 16 percentile. th

The response-classification uncertainty analyses were based on hypothetical classification error

rates. Assumptions of 50% response-classification error for late walking and late talking were worst-case

for those values immediately adjacent to the response decision point value for any given effect. That is,

for late walking, the values of 18 or 20 months for first walking and 24 or 26 months for first talking

were assumed to be equally likely, resulting in misclassification 50% of the time. This would require an

uncertainty in recall of these events of at least 2 months, which is not unlikely in this particular situation. 

The actual classification error was likely to be somewhat less than 50%, particularly as the large number

of observations for late walking at 18 months (22 of the 81 individuals) suggests that 18 months may

have been used as an upper bound in some cases. The response-classification error assumptions for late

walking and late talking were best-case for all other values as no error is assumed. Even with a 25%
D-34

classification error, however, the results of the response-classification uncertainty analysis indicate that

the late walking endpoint was unreliable as a measure of methylmercury toxicity. The exclusion of this

endpoint would not have a very large impact on the combined developmental effects threshold

distribution, increasing the thresholds by about 50%. Although the late talking threshold distribution is

not grossly affected by response-classification error, variability in interpretation of the definition of the

endpoint (first talking) likely would have been greater than that for walking; this uncertainty was not

estimated in this analysis. The neurological effects thresholds were least sensitive to classification error,

assuming that the true error was closer 10% than 20%. The assumption seems reasonable given the much

greater objectivity of the measurement of the effect. Adult paresthesia was the most sensitive to

classification error, showing extreme variability in the threshold estimates with a classification error rate

as low as 5% (all observations). These results suggest that strong conclusions based on the late walking

and adult paresthesia endpoints are unwarranted.

 

Results of the alternate scenarios (Table D-8) show that the primary effect of the correlation

assumptions among the dose conversion input variables was a fairly large reduction in the variance of the

Monte Carlo simulation output. The assumption of correlation of individual susceptibility and half-life

of methylmercury in the blood did not have a marked effect on the simulation except for a 42% reduction

in the median when a strong correlation was assumed (t1⁄2 > 84 days). The latter scenario probably

represented a worst-case situation although no data were found that directly address the magnitude of the

hypothetical correlation. 

The sensitivity analysis indicates that the variables that contribute the most to the dose

conversion simulation variability are the hair:blood ratio (hb), the half-life of methylmercury in the blood

(t1⁄2) and the fraction of absorbed methylmercury found in the blood (f). There is very little that can be

done to reduce the uncertainty in these variables because appropriate data directly applicable to the Iraqi

cohort are not available. These results could be of use in the experimental design and collection of data

for estimates of ingestion levels from hair concentrations in the future. 

2.4.5 Conclusions of Analysis of Uncertainty Around Human Health effects of Methylmercury

A major source of the variability was in the estimation of bootstrap thresholds from the Iraqi

cohort data as evidenced by the 12-20 fold difference in the 5 and 95 percentiles of the bootstrap th th

threshold distributions. The uncertainty arising from limited exposure duration contributed almost as

much, with a 12.5-fold difference in the 5 and 95 percentiles. The corresponding spreads in the dose th th

conversion distributions were 2.4-4.2 fold. Correlations between variables were important with respect

to the variance of the Monte Carlo simulations but were not well-defined by empirical data. Additional

areas of uncertainty remain to be modeled.

Of the developmental endpoints, the neurological effects, which are determined by a battery of

tests and do not depend on subject recall, would seem to be the most objective measure of methylmercury

toxicity. Late walking was not a reliable endpoint because of sensitivity to classification error.

The RfD of 1 x 10 mg/kg-day is very likely well below the threshold for developmental effects -4

but may be above the threshold for exposure duration-adjusted adult paresthesia. Strong conclusions

based on the latter result are not warranted because of the sensitivity of the adult paresthesia threshold to

classification error and the general lack of data addressing the effects of exposure duration.

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D-35

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