TOXICOLOGICAL PROFILE FOR
 

HEXACHLOROBENZENE
 

U.S. DEPARTMENT OF HEALTH AND HUMAN SERVICES
 

Public Health Service
 

Agency for Toxic Substances and Disease Registry
 

September 2002 
HEXACHLOROBENZENE ii 

DISCLAIMER 

The use of company or product name(s) is for identification only and does not imply endorsement by the 

Agency for Toxic Substances and Disease Registry. 
HEXACHLOROBENZENE iii 

UPDATE STATEMENT
 

Toxicological profiles are revised and republished as necessary, but no less than once every three years. 

For information regarding the update status of previously released profiles, contact ATSDR at: 

Agency for Toxic Substances and Disease Registry
 

Division of Toxicology/Toxicology Information Branch
 

1600 Clifton Road NE, E-29
 

Atlanta, Georgia 30333
 

V 

FOREWORD 

This toxicological profile is prepared in accordance with guidelines* developed by the Agency for 

Toxic Substances and Disease Registry (ATSDR) and the Environmental Protection Agency (EPA). The 

original guidelines were published in the FederaZ Register on April 17, 1987. Each profile will be revised 

and republished as necessary. 

The ATSDR toxicological profile succinctly characterizes the toxicologic and adverse health 

effects information for the hazardous substance described therein. Each peer-reviewed profile identifies 

and reviews the key literature that describes a hazardous substance's toxicologic properties. Other 

pertinent literature is also presented, but is described in less detail than the key studies. The profile is not 

intended to be an exhaustive document; however, more comprehensive sources of specialty information 

are referenced. 

The focus of the profiles is on health and toxicologic information; therefore, each toxicological 

profile begins with a public health statement that describes, in nontechnical language, a substance's 

relevant toxicological properties. Following the public health statement is information concerning levels 

of significant human exposure and, where known, significant health effects. The adequacy of 

infomation to determine a substance's health effects is described in a health effects summary. Data 

needs that are of significance to protection of public health are identified by ATSDR and EPA. 

Each profile includes the following: 

The examination, summary, and interpretation of available toxicologic information and 

epidemiologic evaluations on a hazardous substance to ascertain the levels of significant 

human exposure for the substance and the associated acute, subacute, and chronic health 

effects; 

A determination of whether adequate information on the health effects of each substance is 

available or in the process of development to determine levels of exposure that present a 

significant risk to human health of acute, subacute, and chronic health effects; and 

Where appropriate, identification of toxicologic testing needed to identify the types or levels 

of exposure that may present significant risk of adverse health effects in humans. 

The principal audiences for the toxicological profiles are health professionals at the federal, state, 

and local levels; interested private sector organizations and groups; and members of the public. 

This profile reflects ATSDR's assessment of all relevant toxicologic testing and information that 

has been peer-reviewed. Staff of the Centers for Disease Control and Prevention and other federal 

scientists have also reviewed the profile. In addition, this profile has been peer-reviewed by a 

nongovernmental panel and was made available for public review. Final responsibility for the contents 

and views expressed in this toxicological profile resides with ATSDR. 

Administrator 

Agency for Toxic Substances and 

Disease Registry 
HEXACHLOROBENZENE vi 

*Legislative Background 

The toxicological profiles are developed in response to the Superfund Amendments and 

Reauthorization Act (SARA) of 1986 (Public law 99-499) which amended the Comprehensive 

Environmental Response, Compensation, and Liability Act of 1980 (CERCLA or Superfund). This 

public law directed ATSDR to prepared toxicological profiles for hazardous substances most commonly 

found at facilities on the CERCLA National Priorities List and that pose the most significant potential 

threat to human health, as determined by ATSDR and the EPA. The availability of the revised priority 

list of 275 hazardous substances was announced in the Federal Register on November 17, 1997 (62 FR 

61332). For prior versions of the list of substances, see Federal Register notices dated April 29, 1996 (61 

FR 18744); April 17, 1987 (52 FR 12866); October 20, 1988 (53 FR 41280); October 26, 1989 (54 FR 

43619); October 17, 1990 (55 FR 42067); October 17, 1991 (56 FR 52166); October 28, 1992 (57 FR 

48801); and February 28, 1994 (59 FR 9486). Section 104(i)(3) of CERCLA, as amended, directs the 

Administrator of ATSDR to prepare a toxicological profile for each substance on the list. 
HEXACHLOROBENZENE vii 

QUICK REFERENCE FOR HEALTH CARE PROVIDERS 

Toxicological Profiles are a unique compilation of toxicological information on a given hazardous 

substance. Each profile reflects a comprehensive and extensive evaluation, summary, and interpretation of 

available toxicologic and epidemiologic information on a substance. Health care providers treating 

patients potentially exposed to hazardous substances will find the following information helpful for fast 

answers to often-asked questions. 

Primary Chapters/Sections of Interest 

Chapter 1: Public Health Statement: The Public Health Statement can be a useful tool for educating 

patients about possible exposure to a hazardous substance. It explains a substance’s relevant 

toxicologic properties in a nontechnical, question-and-answer format, and it includes a review of 

the general health effects observed following exposure. 

Chapter 2: Relevance to Public Health: The Relevance to Public Health Section evaluates, interprets, 

and assesses the significance of toxicity data to human health. 

Chapter 3: Health Effects: Specific health effects of a given hazardous compound are reported by type 

of health effect (death, systemic, immunologic, reproductive), by route of exposure, and by length 

of exposure (acute, intermediate, and chronic). In addition, both human and animal studies are 

reported in this section. 

NOTE: Not all health effects reported in this section are necessarily observed in 

the clinical setting. Please refer to the Public Health Statement to identify 

general health effects observed following exposure. 

Pediatrics: Four new sections have been added to each Toxicological Profile to address child health 

issues: 

Section 1.6 How Can (Chemical X) Affect Children?
 

Section 1.7 How Can Families Reduce the Risk of Exposure to (Chemical X)?
 

Section 3.7 Children’s Susceptibility
 

Section 6.6 Exposures of Children
 

Other Sections of Interest: 

Section 3.8 Biomarkers of Exposure and Effect 

Section 3.11 Methods for Reducing Toxic Effects 

ATSDR Information Center 

Phone: 1-888-42-ATSDR or (404) 498-0110 Fax: (404) 498-0057 

E-mail: atsdric@cdc.gov Internet: http://www.atsdr.cdc.gov 

The following additional material can be ordered through the ATSDR Information Center: 

Case Studies in Environmental Medicine: Taking an Exposure History—The importance of taking an 

exposure history and how to conduct one are described, and an example of a thorough exposure 

history is provided. Other case studies of interest include Reproductive and Developmental 

Hazards; Skin Lesions and Environmental Exposures; Cholinesterase-Inhibiting Pesticide 

Toxicity; and numerous chemical-specific case studies. 
HEXACHLOROBENZENE viii 

Managing Hazardous Materials Incidents is a three-volume set of recommendations for on-scene 

(prehospital) and hospital medical management of patients exposed during a hazardous materials incident. 

Volumes I and II are planning guides to assist first responders and hospital emergency department 

personnel in planning for incidents that involve hazardous materials. Volume III—Medical Management 

Guidelines for Acute Chemical Exposures—is a guide for health care professionals treating patients 

exposed to hazardous materials. 

Fact Sheets (ToxFAQs) provide answers to frequently asked questions about toxic substances. 

Other Agencies and Organizations 

The National Center for Environmental Health (NCEH) focuses on preventing or controlling disease, 

injury, and disability related to the interactions between people and their environment outside the 

workplace. Contact: NCEH, Mailstop F-29, 4770 Buford Highway, NE, Atlanta, GA 30341­

3724 • Phone: 770-488-7000 • FAX: 770-488-7015. 

The National Institute for Occupational Safety and Health (NIOSH) conducts research on occupational 

diseases and injuries, responds to requests for assistance by investigating problems of health and 

safety in the workplace, recommends standards to the Occupational Safety and Health 

Administration (OSHA) and the Mine Safety and Health Administration (MSHA), and trains 

professionals in occupational safety and health. Contact: NIOSH, 200 Independence Avenue, 

SW, Washington, DC 20201 • Phone: 800-356-4674 or NIOSH Technical Information Branch, 

Robert A. Taft Laboratory, Mailstop C-19, 4676 Columbia Parkway, Cincinnati, OH 45226-1998 

• Phone: 800-35-NIOSH. 

The National Institute of Environmental Health Sciences (NIEHS) is the principal federal agency for 

biomedical research on the effects of chemical, physical, and biologic environmental agents on 

human health and well-being. Contact: NIEHS, PO Box 12233, 104 T.W. Alexander Drive, 

Research Triangle Park, NC 27709 • Phone: 919-541-3212. 

Referrals 

The Association of Occupational and Environmental Clinics (AOEC) has developed a network of clinics 

in the United States to provide expertise in occupational and environmental issues. Contact: 

AOEC, 1010 Vermont Avenue, NW, #513, Washington, DC 20005 • Phone: 202-347-4976 • 

FAX: 202-347-4950 • e-mail: AOEC@AOEC.ORG • Web Page: http://www.aoec.org/. 

The American College of Occupational and Environmental Medicine (ACOEM) is an association of 

physicians and other health care providers specializing in the field of occupational and 

environmental medicine. Contact: ACOEM, 55 West Seegers Road, Arlington Heights, IL 

60005 • Phone: 847-818-1800 • FAX: 847-818-9266. 
HEXACHLOROBENZENE ix 

CONTRIBUTORS 

CHEMICAL MANAGER(S)/AUTHORS(S): 

Jessilynn Taylor, M.S.
 

ATSDR, Division of Toxicology, Atlanta, Georgia
 

Jewell Wilson, Ph.D.
 

ATSDR, Division of Toxicology, Atlanta, Georgia
 

Marc Odin, M.S., D.A.B.T.
 

Syracuse Research Corporation, North Syracuse, New York
 

Andrew J. McDougal, Ph.D.
 

Syracuse Research Corporation, North Syracuse, New York
 

Richard J. Amata, Ph.D.
 

Syracuse Research Corporation, North Syracuse, New York
 

Philip Howard, Ph.D.
 

Syracuse Research Corporation, North Syracuse, New York
 

Farzana Dalal, B.S.
 

Syracuse Research Corporation, North Syracuse, New York
 

THE PROFILE HAS UNDERGONE THE FOLLOWING ATSDR INTERNAL REVIEWS: 

1.	 Health Effects Review. The Health Effects Review Committee examines the health effects 

chapter of each profile for consistency and accuracy in interpreting health effects and classifying 

end points. 

2.	 Minimal Risk Level Review. The Minimal Risk Level Workgroup considers issues relevant to 

substance-specific minimal risk levels (MRLs), reviews the health effects database of each 

profile, and makes recommendations for derivation of MRLs. 

3.	 Data Needs Review. The Research Implementation Branch reviews data needs sections to assure 

consistency across profiles and adherence to instructions in the Guidance. 

HEXACHLOROBENZENE xi 

PEER REVIEW
 

A peer review panel was assembled for hexachlorobenzene. The panel consisted of the following 

members: 

1.	 Dr. Syed M. GhiasUddin, Indiana Department of Environmental Management, 1815 Beckenbauer 

Lane, Indianapolis, IN 46214; 

2.	 Dr. Kannan Krishnan, University of Montreal, Faculty of Medicine, 2375 Chemin de la Cote SteCatherine, Room 4105, Montreal, QC H3T 1A8; and 

3.	 Dr. Clint Skinner, Skinner Associates, Consultants in Toxicology, 3985 Shooting Star Road, 

Creston, CA 93432. 

These experts collectively have knowledge of hexachlorobenzene's physical and chemical properties, 

toxicokinetics, key health end points, mechanisms of action, human and animal exposure, and 

quantification of risk to humans. All reviewers were selected in conformity with the conditions for peer 

review specified in Section 104(I)(13) of the Comprehensive Environmental Response, Compensation, 

and Liability Act, as amended. 

Scientists from the Agency for Toxic Substances and Disease Registry (ATSDR) have reviewed the peer 

reviewers' comments and determined which comments will be included in the profile. A listing of the 

peer reviewers' comments not incorporated in the profile, with a brief explanation of the rationale for their 

exclusion, exists as part of the administrative record for this compound. A list of databases reviewed and 

a list of unpublished documents cited are also included in the administrative record. 

The citation of the peer review panel should not be understood to imply its approval of the profile's final 

content. The responsibility for the content of this profile lies with the ATSDR. 

HEXACHLOROBENZENE xiii 

CONTENTS
 

FOREWORD ....................................................................... v 
 

QUICK REFERENCE FOR HEALTH CARE PROVIDERS ................................. vii 
 

CONTRIBUTORS ................................................................... ix 
 

PEER REVIEW ..................................................................... xi 
 

LIST OF FIGURES ................................................................ xvii 
 

LIST OF TABLES .................................................................. xix 
 

1. PUBLIC HEALTH STATEMENT .................................................... 1 
 

1.1 	 WHAT IS HEXACHLOROBENZENE? .......................................... 1 
 

1.2	 WHAT HAPPENS TO HEXACHLOROBENZENE WHEN IT ENTERS THE
 

ENVIRONMENT? ........................................................... 2 
 

1.3 	 HOW MIGHT I BE EXPOSED TO HEXACHLOROBENZENE? ...................... 2 
 

1.4	 HOW CAN HEXACHLOROBENZENE ENTER AND LEAVE MY BODY? ............ 3 
 

1.5	 HOW CAN HEXACHLOROBENZENE AFFECT MY HEALTH? ..................... 4 
 

1.6	 HOW CAN HEXACHLOROBENZENE AFFECT CHILDREN? ...................... 5 
 

1.7	 HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO
 

HEXACHLOROBENZENE? ................................................... 7 
 

1.8	 IS THERE A MEDICAL TEST TO DETERMINE WHETHER I HAVE BEEN EXPOSED 

TO HEXACHLOROBENZENE? ............................................... 8 
 

1.9	 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO
 

PROTECT HUMAN HEALTH? ................................................ 9 
 

1.10 WHERE CAN I GET MORE INFORMATION? .................................. 10 
 

2. RELEVANCE TO PUBLIC HEALTH ............................................... 11 
 

2.1	 BACKGROUND AND ENVIRONMENTAL EXPOSURES TO HEXACHLOROBENZENE
 

IN THE UNITED STATES ................................................... 11 
 

2.2	 SUMMARY OF HEALTH EFFECTS ........................................... 12 
 

2.3	 MINIMAL RISK LEVELS (MRLs) ............................................ 16 
 

3. HEALTH EFFECTS .............................................................. 23 
 

3.1	 INTRODUCTION .......................................................... 23 
 

3.2	 DISCUSSION OF HEALTH EFFECTS BY ROUTE OF EXPOSURE ................. 23 
 

3.2.1 Inhalation Exposure .................................................. 25 
 

3.2.1.1 Death ..................................................... 25 
 

3.2.1.2 Systemic Effects ............................................ 25 
 

3.2.1.3 Immunological and Lymphoreticular Effects ...................... 33 
 

3.2.1.4 Neurological Effects ......................................... 34 
 

3.2.1.5 Reproductive Effects ......................................... 34 
 

3.2.1.6 Developmental Effects ....................................... 35 
 

3.2.1.7 Cancer .................................................... 35 
 

3.2.2 Oral Exposure ...................................................... 36 
 
HEXACHLOROBENZENE xiv 

3.2.2.1 Death ..................................................... 36 
 

3.2.2.2 Systemic Effects ............................................ 38 
 

3.2.2.3 Immunological and Lymphoreticular Effects ...................... 95
 

3.2.2.4 Neurological Effects ........................................ . 99
 

3.2.2.5 Reproductive Effects ........................................ 102
 

3.2.2.6 Developmental Effects ...................................... 107
 

3.2.2.7 Cancer ................................................... 111
 

3.2.3 Dermal Exposure ................................................... 115
 

3.2.3.1 Death .................................................... 115
 

3.2.3.2 Systemic Effects ........................................... 115
 

3.2.3.3 Immunological and Lymphoreticular Effects ..................... 115
 

3.2.3.4 Neurological Effects ........................................ 115
 

3.2.3.5 Reproductive Effects ........................................ 115
 

3.2.3.6 Developmental Effects ...................................... 115
 

3.2.3.7 Cancer ................................................... 116
 

3.3 GENOTOXICITY ......................................................... 116
 

3.4 TOXICOKINETICS ........................................................ 118
 

3.4.1 Absorption ........................................................ 123
 

3.4.1.1 Inhalation Exposure ......................................... 123
 

3.4.1.2 Oral Exposure ............................................. 125
 

3.4.1.3 Dermal Exposure ........................................... 128
 

3.4.2	 Distribution ....................................................... 129
 

3.4.2.1 Inhalation Exposure ......................................... 129
 

3.4.2.2 Oral Exposure ............................................. 130
 

3.4.2.3 Dermal Exposure ........................................... 135
 

3.4.3	 Metabolism ........................................................ 135
 

3.4.3.1 Inhalation Exposure ......................................... 135
 

3.4.3.2 Oral Exposure ............................................. 135
 

3.4.3.3 Dermal Exposure ........................................... 140
 

3.4.4	 Elimination and Excretion ............................................ 140
 

3.4.4.1 Inhalation Exposure ......................................... 140
 

3.4.4.2 Oral Exposure ............................................. 141
 

3.4.4.3 Dermal Exposure ........................................... 143
 

3.4.4.4 Other Routes of Exposure .................................... 143
 

3.4.5	 Physiologically Based Pharmacokinetic (PBPK)/Pharmacodynamic (PD) 

Models ........................................................... 143
 

3.4.5.1 Summary of PBPK Models ................................... 145
 

3.4.5.2 Hexachlorobenzene PBPK Model Comparison ................... 145
 

3.4.5.3 Discussion of Models ....................................... 147
 

3.5 MECHANISMS OF ACTION ................................................ 149
 

3.5.1	 Pharmacokinetic Mechanisms ......................................... 149
 

3.5.2	 Mechanisms of Toxicity .............................................. 152
 

3.5.3	 Animal-to-Human Extrapolations ...................................... 155
 

3.6 TOXICITIES MEDIATED THROUGH THE NEUROENDOCRINE AXIS ............ 155
 

3.7 CHILDREN’S SUSCEPTIBILITY ............................................ 157
 

3.8 BIOMARKERS OF EXPOSURE AND EFFECT ................................. 160
 

3.8.1 Biomarkers Used to Identify or Quantify Exposure to Hexachlorobenzene ...... 161
 

3.8.2 Biomarkers Used to Characterize Effects Caused by Hexachlorobenzene ....... 162
 

3.9 INTERACTIONS WITH OTHER CHEMICALS ................................. 163
 

3.10 POPULATIONS THAT ARE UNUSUALLY SUSCEPTIBLE ...................... 167
 

3.11 METHODS FOR REDUCING TOXIC EFFECTS ................................ 168
 
HEXACHLOROBENZENE xv 

3.11.1 Reducing Peak Absorption Following Exposure ........................... 168
 

3.11.2 Reducing Body Burden .............................................. 169
 

3.11.3 Interfering with the Mechanism of Action for Toxic Effects .................. 170
 

3.12 ADEQUACY OF THE DATABASE ........................................... 171
 

3.12.1 Existing Information on Health Effects of Hexachlorobenzene ............... 172
 

3.12.2 Identification of Data Needs .......................................... 172
 

3.12.3 Ongoing Studies .................................................... 186
 

4. CHEMICAL AND PHYSICAL INFORMATION ...................................... 187
 

4.1 CHEMICAL IDENTITY .................................................... 187
 

4.2 PHYSICAL AND CHEMICAL PROPERTIES ................................... 187
 

5. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL ............................ 191
 

5.1 PRODUCTION ........................................................... 191
 

5.2 IMPORT/EXPORT ........................................................ 192
 

5.3 USE ..................................................................... 194
 

5.4 DISPOSAL ............................................................... 194
 

6. POTENTIAL FOR HUMAN EXPOSURE ........................................... 197
 

6.1 OVERVIEW .............................................................. 197
 

6.2 RELEASES TO THE ENVIRONMENT ........................................ 200
 

6.2.1 Air .............................................................. 201
 

6.2.2 Water ............................................................ 204
 

6.2.3 Soil .............................................................. 206
 

6.3 ENVIRONMENTAL FATE .................................................. 208
 

6.3.1 Transport and Partitioning ............................................ 208
 

6.3.2 Transformation and Degradation ....................................... 213
 

6.3.2.1 Air ...................................................... 213
 

6.3.2.2 Water .................................................... 214
 

6.3.2.3 Sediment and Soil .......................................... 215
 

6.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT ............... 216
 

6.4.1 Air .............................................................. 217
 

6.4.2 Water ............................................................ 219
 

6.4.3 Sediment and Soil .................................................. 220
 

6.4.4 Other Environmental Media ........................................... 222
 

6.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSURE .................. 232
 

6.6 EXPOSURES OF CHILDREN ............................................... 236
 

6.7 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES ...................... 240
 

6.8 ADEQUACY OF THE DATABASE ........................................... 242
 

6.8.1 Identification of Data Needs .......................................... 243
 

6.8.2 Ongoing Studies .................................................... 248
 

7. ANALYTICAL METHODS ....................................................... 250
 

7.1 BIOLOGICAL MATERIALS ................................................ 250
 

7.2 ENVIRONMENTAL SAMPLES .............................................. 254
 

7.3 ADEQUACY OF THE DATABASE ........................................... 261
 

7.3.1 Identification of Data Needs .......................................... 262
 

7.3.2 Ongoing Studies .................................................... 264
 

8. REGULATIONS AND ADVISORIES .............................................. 266
 
HEXACHLOROBENZENE xvi 

9. REFERENCES ................................................................. 274
 

10. GLOSSARY .................................................................. 346
 

APPENDICES
 

A. ATSDR MINIMAL RISK LEVELS AND WORKSHEETS ......................... A-1 
 

B. USER’S GUIDE ........................................................... B-1 
 

C. ACRONYMS, ABBREVIATIONS, AND SYMBOLS ............................. C-1 
 

D. INDEX .................................................................. D-1 
 
HEXACHLOROBENZENE xvii 

LIST OF FIGURES
 

3-1. Levels of Significant Exposure to Hexachlorobenzene - Inhalation ........................ 30 
 

3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral ........................... 70 
 

3-3. Metabolism and Urinary Metabolites of Hexachlorobenzene ............................ 139
 

3-4. Conceptual Representation of a Physiologically Based Pharmacokinetic (PBPK) Model for a 

Hypothetical Chemical Substance ................................................ 146
 

3-5. Existing Information on Health Effects of Hexachlorobenzene .......................... 174
 

6-1. Frequency of NPL Sites with Hexachlorobenzene Contamination ........................ 198
 

HEXACHLOROBENZENE xix 

LIST OF TABLES 

3-1. Levels of Significant Exposure to Hexachlorobenzene - Inhalation ....................... 27 
 

3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral ............................. 39 
 

3-3. Genotoxicity of Hexachlorobenzene In Vivo ......................................... 119
 

3-4. Genotoxicity of Hexachlorobenzene In Vitro ........................................ 120
 

4-1. Chemical Identity of Hexachlorobenzene ........................................... 188
 

4-2. Physical and Chemical Properties of Hexachlorobenzene ............................... 189
 

5-1. Facilities that Produce, Process, or Use Hexachlorobenzene ............................ 193
 

6-1. U.S. Hexachlorobenzene Emissions Summary, Mid 1990s, Part A ....................... 202
 

6-1. U.S. Hexachlorobenzene Emissions Summary, Mid 1990s, Part B ....................... 203
 

6-2. Releases to the Environment From Facilities That Produce, Process, or Use 

Hexachlorobenzene ............................................................ 205
 

6-3. Sites with the Five Highest Concentrations of Hexachlorobenzene in Fish ................. 225
 

6-4. Residue Concentrations (ng/g) in Snapping Turtle Eggs from Five Locations Collected in
 

1981–1991 .................................................................. 228
 

6-5. Mean Levels of Hexachlorobenzene in Breast Milk ................................... 238
 

6-6. Fish Consumption Advisories .................................................... 241
 

7-1. Analytical Methods for Determining Hexachlorobenzene in Biological Samples ............ 251
 

7-2. Analytical Methods for Determining Hexachlorobenzene in Environmental Samples ......... 255
 

7-3. Analytical Methods for Determining Biomarkers of Hexachlorobenzene .................. 263
 

8-1. Regulations and Guidelines Applicable to Hexachlorobenzene .......................... 268
 

HEXACHLOROBENZENE 1 

1. PUBLIC HEALTH STATEMENT
 

This public health statement tells you about hexachlorobenzene and the effects of exposure. 

The Environmental Protection Agency (EPA) identifies the most serious hazardous waste sites in 

the nation. These sites make up the National Priorities List (NPL) and are the sites targeted for 

long-term federal cleanup activities. Hexachlorobenzene has been found in at least 107 of the 

1,613 current or former NPL sites. However, the total number of NPL sites evaluated for this 

substance is not known. As more sites are evaluated, the sites at which hexachlorobenzene is 

found may increase. This information is important because exposure to this substance may harm 

you and because these sites may be sources of exposure. 

When a substance is released from a large area, such as an industrial plant, or from a container, 

such as a drum or bottle, it enters the environment. This release does not always lead to 

exposure. You are exposed to a substance only when you come in contact with it. You may be 

exposed by breathing, eating, or drinking the substance, or by skin contact. 

If you are exposed to hexachlorobenzene, many factors determine whether you’ll be harmed. 

These factors include the dose (how much), the duration (how long), and how you come in 

contact with it. You must also consider the other chemicals you’re exposed to and your age, sex, 

diet, family traits, lifestyle, and state of health. 

1.1 WHAT IS HEXACHLOROBENZENE? 

Hexachlorobenzene is a white crystalline solid. This compound does not occur naturally. It is 

formed as a by-product during the manufacture of chemicals used as solvents (substances used to 

dissolve other substances), other chlorine-containing compounds, and pesticides. Small amounts 

of hexachlorobenzene can also be produced during combustion processes such as burning of city 

wastes. It may also be produced as a by-product in waste streams of chlor-alkali and woodpreserving plants. Hexachlorobenzene was widely used as a pesticide until 1965. It was also 

used to make fireworks, ammunition, and synthetic rubber. Currently, the substance is not used 
HEXACHLOROBENZENE 2 

1. PUBLIC HEALTH STATEMENT 

commercially in the United States. For more information on the physical and chemical 

properties of this substance, see Chapter 4. 

1.2	 WHAT HAPPENS TO HEXACHLOROBENZENE WHEN IT ENTERS THE 

ENVIRONMENT? 

Hexachlorobenzene tends to remain in the environment for a long time. If it is released to the 

soil, it has a half-life of 3–6 years. This means that half of the total amount will disappear after 

3–6 years, half of the remaining amount will disappear in another 3–6 years, and this process 

will continue each 3–6 years thereafter. If it is released to surface waters such as lakes, rivers, 

and streams, the half-life is 2.7–5.7 years, and if it is released to groundwater, the half-life is 

5.3–11.4 years. Since hexachlorobenzene does not dissolve in water very well, most of it will 

remain in particles on the bottom of lakes, rivers, or streams. The evaporation of hexachlorobenzene into the air is not significant under ordinary conditions. Once in the air, it can be 

carried over wide geographic areas. Its half-life in air ranges from 0.63 to 6.28 years. For more 

information on the releases, occurrence, and movement of this substance, see Chapters 4, 5, 

and 6. 

1.3	 HOW MIGHT I BE EXPOSED TO HEXACHLOROBENZENE? 

You may be exposed to hexachlorobenzene if you live near an industrial site where it is 

produced as an unintentional by-product or as a minor part of another chemical product. You 

may also be exposed if you live near a hazardous waste site where hexachlorobenzene has been 

discarded. At these sites, hexachlorobenzene may be carried in the air on dust particles. The 

levels of it in the air around these sites are not known. The amount you might breathe in urban 

air is very small, about 0.00003 parts per million (ppm). If you work in an industry that uses or 

produces it unintentionally as a by-product, you may also breathe in hexachlorobenzene particles 

or dust that carries it, and it may touch your skin. In workplaces where hexachlorobenzene is 

used or disposed, an estimated 1–1.4 ppm of this substance has been measured in the air. You 

can also be exposed to it through contact with soil or dust particles contaminated with 

hexachlorobenzene, or by breathing air contaminated by industrial releases into the environment. 
HEXACHLOROBENZENE 3 

1. PUBLIC HEALTH STATEMENT 

Because it does not dissolve easily in water, this substance is usually not present in high 

concentrations in drinking water. Based on a national survey, it was detected at very low levels 

in surface water and groundwater systems. Drinking water in three cities contained only trace 

amounts of hexachlorobenzene. Therefore, exposure to this substance through drinking water is 

limited. 

Eating foods such as shellfish, fish, and certain vegetables can also expose people to 

hexachlorobenzene. Additionally, you can be exposed to hexachlorobenzene by eating and 

drinking foods and liquids, such as milk, other dairy products, meat, and poultry, if the animals 

from which these products are obtained have been exposed to it through their feed or other 

sources of contamination. Additionally, fat and oil in food may increase the amount of 

hexachlorobenzene that enters the body from food. 

Low levels of hexachlorobenzene have been found in the fatty tissues of almost all people tested. 

These amounts are most likely the result of low levels in food. An estimated average yearly 

uptake of 1 microgram per kilogram (μg/kg) of body weight has been calculated for exposure to 

contaminated food. Exposure by breathing contaminated air is estimated to be 0.01 μg/kg/year. 

This is 100 times less than the exposure from eating contaminated foods. Drinking contaminated 

water is estimated to contribute only very small amounts (0.00085 μg/kg/year). For more 

information, see Chapter 6. 

1.4 HOW CAN HEXACHLOROBENZENE ENTER AND LEAVE MY BODY? 

Hexachlorobenzene can enter your body when you eat food contaminated with it, when you 

breathe particles of it in the air, or when it gets on your skin. After it enters your body, it rapidly 

spreads through your blood to many tissues in the body, especially to fat. This probably happens 

within a few hours. Based on the results of a survey of this substance in people's tissues, it will 

remain in your body, especially in fat, for years. A large portion of hexachlorobenzene in the fat 

of a mother can be transferred to her baby in breast milk. During pregnancy, this substance can 

transfer to the unborn child through the mother's blood. Most of it leaves your body in the feces; 

smaller amounts are found in urine. For more information, see Chapter 3. 
HEXACHLOROBENZENE 4 

1. PUBLIC HEALTH STATEMENT 

1.5 HOW CAN HEXACHLOROBENZENE AFFECT MY HEALTH? 

To protect the public from the harmful effects of toxic chemicals and to find ways to treat people 

who have been harmed, scientists use many tests. 

One way to see if a chemical will hurt people is to learn how the chemical is absorbed, used, and 

released by the body; for some chemicals, animal testing may be necessary. Animal testing may 

also be used to identify health effects such as cancer or birth defects. Without laboratory 

animals, scientists would lose a basic method to get information needed to make wise decisions 

to protect public health. Scientists have the responsibility to treat research animals with care and 

compassion. Laws today protect the welfare of research animals, and scientists must comply 

with strict animal care guidelines. 

A study of people in Turkey who, over a period of several years, ate grain that was accidentally 

contaminated with hexachlorobenzene, showed a high death rate in young children of mothers 

who ate it and in young children who ate it themselves. Nursing infants may be in particular 

danger from the transfer of hexachlorobenzene through breast milk if their mothers have been 

exposed. Unborn children may also be at particular risk because of the transfer of hexachlorobenzene through the mother's blood during pregnancy. This has been confirmed by experiments 

in animals. 

People in Turkey who, over a long time, ate grain that was accidentally contaminated with 

hexachlorobenzene suffered from a liver disease called porphyria cutanea tarda. The main effect 

of porphyria is slowed or stopped formation of heme, the oxygen-carrying part of the 

hemoglobin molecule found in red blood cells and an important chemical in the body. Porphyria 

is identified by elevation of heme precursors called porphyrins in the blood, urine, and stool. 

This disease can cause red-colored urine, skin sores, change in skin color, arthritis, and problems 

of the liver, nervous system, and stomach. Studies in animals also show that eating hexachlorobenzene for a long time can harm mostly the liver, thyroid, and nervous systems; the studies in 

animals also show that eating hexachlorobenzene for months or years can damage bones, 
HEXACHLOROBENZENE 5 

1. PUBLIC HEALTH STATEMENT 

kidneys, and blood, and the immune, endocrine (hormone-releasing), and nervous systems. 

Unborn children and young children may be more sensitive to these effects than adults. 

Studies in animals also suggest that eating this substance for months or years can cause cancer of 

the liver, kidney, and thyroid, but there is no strong evidence that the substance causes cancer in 

people. The U.S. Department of Health and Human Services (DHHS) has determined that the 

substance may reasonably be anticipated to be a human carcinogen. The International Agency 

for Research on Cancer (IARC) has determined that hexachlorobenzene is possibly carcinogenic 

to humans. EPA has determined that hexachlorobenzene is a probable human carcinogen. See 

Chapter 3 for more information. 

There is no information on levels of hexachlorobenzene in air that affect people. No information 

is available on the taste or lowest levels at which you can taste or smell it in air or water. 

Therefore, people may not know when hexachlorobenzene is present in air or water. Studies 

conducted with rats indicate that breathing hexachlorobenzene may harm the immune system. 

The health effects from getting it on your skin are not known. 

For more information on the levels of this substance that have caused health effects in people 

and animals, see Chapters 2 and 3. 

1.6 HOW CAN HEXACHLOROBENZENE AFFECT CHILDREN? 

This section discusses potential health effects from exposures during the period from conception 

to maturity at 18 years of age in humans. 

Infants and young children appeared to be especially sensitive to the effects of hexachlorobenzene in the Turkish grain poisoning epidemic. Breast-fed infants of mothers known to have 

ingested the contaminated bread developed a disease known as pembe yara or "pink sore." The 

disease was named for the skin lesions that were produced. Other symptoms were weakness and 

convulsions. Many of the sickened infants died from this disease. Young children beyond 
HEXACHLOROBENZENE 6 

1. PUBLIC HEALTH STATEMENT 

2 years of age did not get pink sore, but they did develop numerous skin, nervous system, and 

bone abnormalities later in life. 

One study found higher levels of hexachlorobenzene in mother’s milk in babies who had ear 

infections than in the milk of babies without ear infections. Another study found higher levels of 

hexachlorobenzene in the fat of boys with a specific type of birth defect, undescended testis, than 

in the fat of normal boys. We do not know if hexachlorobenzene caused the infection or the 

birth defect. 

Animal studies support the suggestion that young animals exposed to hexachlorobenzene before 

and soon after birth are especially sensitive to hexachlorobenzene. Effects on the liver, nervous 

system, and immune function occurred at lower doses in the young developing animals than in 

adults. Animal studies also showed that hexachlorobenzene has effects on various endocrine 

organs, including the thyroid gland (hypothyroidism), parathyroid gland (hyperparathyroidism), 

adrenal gland, and ovaries. These tissues produce hormones that are important to normal growth 

and development of the organism. 

You can find more information about how hexachlorobenzene can affect children in Sections 3.7 

and 6.6. 
HEXACHLOROBENZENE 7 

1. PUBLIC HEALTH STATEMENT 

1.7 HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO
 

HEXACHLOROBENZENE? 

If your doctor finds that you have been exposed to significant amounts of hexachlorobenzene, 

ask whether your children might also be exposed. Your doctor might need to ask your state 

health department to investigate. 

The primary way most people are exposed is through food. Fatty food may be higher in 

hexachlorobenzene than less fatty food. Additionally, when hexachlorobenzene is present in 

food, more may be absorbed when the food is fatty than when the food is less fatty. Therefore, 

eating less fatty food may reduce the risk of exposure to hexachlorobenzene. 

If you live near an industrial site where hexachlorobenzene was produced or is produced as an 

unintentional by-product or you live near a hazardous waste site where it has been discarded, 

there may be high levels of hexachlorobenzene in the water and soil. Substituting cleaner 

sources of water and limiting contact with soil (for example, through use of a dense ground cover 

or thick lawn) would reduce family exposure to hexachlorobenzene. Produce grown in 

contaminated soil should not be eaten. By paying careful attention to dust and dirt control in the 

home (air filters, frequent cleaning), you can reduce family exposure to contaminated dirt. Some 

children eat a lot of dirt. You should prevent your children from eating dirt. You should 

discourage your children from putting objects in their mouths. Make sure that they wash their 

hands frequently and before eating. Discourage your children from putting their hands in their 

mouths or from other hand-to-mouth activity. 

It is sometimes possible to carry hexachlorobenzene from work on your clothing, skin, hair, 

tools, or other objects removed from the workplace. This is particularly likely if you work in the 

chemical or pesticide industries. You may contaminate your car, home, or other locations 

outside work where children might be exposed to hexachlorobenzene. You should know about 

this possibility if you work where hexachlorobenzene exposure may occur. 
HEXACHLOROBENZENE 8 

1. PUBLIC HEALTH STATEMENT 

Your occupational health and safety officer at work can and should tell you whether chemicals 

you work with are dangerous and likely to be carried home on your clothes, body, or tools and 

whether you should be showering and changing clothes before you leave work, storing your 

street clothes in a separate area of the workplace, or laundering your work clothes at home 

separately from other clothes. Material safety data sheets (MSDS) for many chemicals used 

should be found at your place of work, as required by the Occupational Safety and Health 

Administration (OSHA) in the U.S. Department of Labor. MSDS information should include 

chemical names and hazardous ingredients, and important properties, such as fire and explosion 

data, potential health effects, how you get the chemical(s) in your body, how to properly handle 

the materials, and what to do in the case of emergencies. Your employer is legally responsible 

for providing a safe workplace and should freely answer your questions about hazardous 

chemicals. Your state OSHA-approved occupational safety and health program or OSHA itself 

can answer any further questions and help your employer identify and correct problems with 

hazardous substances. Your state OSHA-approved occupational safety and health program or 

OSHA will listen to your formal complaints about workplace health hazards and inspect your 

workplace when necessary. Employees have a right to seek safety and health on the job without 

fear of punishment. 

1.8	 IS THERE A MEDICAL TEST TO DETERMINE WHETHER I HAVE BEEN 

EXPOSED TO HEXACHLOROBENZENE? 

Blood, breast milk, urine, and feces may be tested to determine if you have ever been exposed to 

hexachlorobenzene. These tests are not usually available at a doctor's office because they 

require special equipment found at county, state, university, and independent analytical 

laboratories. Because this compound can collect and remain in human fat for several years, the 

test for this substance in breast milk can tell you only that you have been exposed, but not when 

or to how much. The blood, urine, and feces tests can indicate more recent exposure. However, 

the levels measured cannot be used to predict what health effects might occur in people. Blood, 

urine, and feces can also be monitored for porphyrins. High porphyrin levels indicate slowed 

formation of heme, which is a major effect of hexachlorobenzene in the body. The usefulness of 
HEXACHLOROBENZENE 9 

1. PUBLIC HEALTH STATEMENT 

this test as a sign of hexachlorobenzene exposure is limited, however, because there are many 

other potential causes of high porphyrin levels. For more information, see Chapters 3 and 7. 

1.9 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO 

PROTECT HUMAN HEALTH? 

The federal government develops regulations and recommendations to protect public health. 

Regulations can be enforced by law. Federal agencies that develop regulations for toxic 

substances include the Environmental Protection Agency (EPA), the Occupational Safety and 

Health Administration (OSHA), and the Food and Drug Administration (FDA). 

Recommendations provide valuable guidelines to protect public health but cannot be enforced by 

law. Federal organizations that develop recommendations for toxic substances include the 

Agency for Toxic Substances and Disease Registry (ATSDR) and the National Institute for 

Occupational Safety and Health (NIOSH). 

Regulations and recommendations can be expressed in not-to-exceed levels in air, water, soil, or 

food that are usually based on levels that affect animals; then they are adjusted to help protect 

people. Sometimes these not-to-exceed levels differ among federal organizations because of 

different exposure times (an 8-hour workday or a 24-hour day), the use of different animal 

studies, or other factors. 

Recommendations and regulations are also periodically updated as more information becomes 

available. For the most current information, check with the federal agency or organization that 

provides it. Some regulations and recommendations for hexachlorobenzene include the 

following: 

EPA has developed advisories to protect people from the potential health effects of exposure to 

hexachlorobenzene in drinking water. EPA has proposed that drinking water should not contain 

more than 0.05 parts of hexachlorobenzene per million parts of water (ppm) in water that 

children drink, and should not exceed 0.2 ppm in water that adults drink for longer periods 

(approximately 7 years). No other health standards or recommendations exist for this substance. 
HEXACHLOROBENZENE 10 

1. PUBLIC HEALTH STATEMENT 

You can find more information on regulations and guidelines that apply to hexachlorobenzene in 

Chapter 8. 

1.10 WHERE CAN I GET MORE INFORMATION? 

If you have any more questions or concerns, please contact your community or state health or 

environmental quality department or 

Agency for Toxic Substances and Disease Registry
 

Division of Toxicology
 

1600 Clifton Road NE, Mailstop E-29
 

Atlanta, GA 30333
 

Web site: http://www.atsdr.cdc.gov
 

* Information line and technical assistance 

Phone: 1-888-42-ATSDR (1-888-422-8737)
 

Fax: 1-404-498-0057
 

ATSDR can also tell you the location of occupational and environmental health clinics. These 

clinics specialize in recognizing, evaluating, and treating illnesses resulting from exposure to 

hazardous substances. 

* To order toxicological profiles, contact 

National Technical Information Service
 

5285 Port Royal Road
 

Springfield, VA 22161
 

Phone: (800) 553-6847 or (703) 605-6000
 
HEXACHLOROBENZENE 11 

2. RELEVANCE TO PUBLIC HEALTH
 

2.1	 BACKGROUND AND ENVIRONMENTAL EXPOSURES TO HEXACHLOROBENZENE 

IN THE UNITED STATES 

Hexachlorobenzene is a fully chlorinated hydrocarbon industrial chemical that is practically insoluble in 

water, but is very soluble in fat, oils, and organic solvents. Hexachlorobenzene is one of the most 

persistent environmental pollutants, and bioaccumulates in the environment, in animals, and in humans. 

It is not currently manufactured as a commercial product in the United States, and virtually all 

commercial production ended in the late 1970s. However, some hexachlorobenzene is produced as a byproduct or impurity in the manufacture of chlorinated solvents and other chlorinated compounds (e.g., 

carbon tetrachloride, chlorophenols, mirex, tetrachloroethylene, trichlorobenzenes, trichloroethylene, 

trichlorotoluenes, and vinyl chloride), including several pesticides currently in use (pentachloronitrobenzene, chlorothalonil, Dacthal®, picloram, pentachlorophenol, atrazine, simazine, and lindane). It 

is estimated that 3,500–11,500 kg of hexachlorobenzene were inadvertently produced in the manufacture 

of chlorinated solvents in 1984. There are no current commercial uses of hexachlorobenzene in the 

United States, although hexachlorobenzene was used as a fungicide on the seeds of onions, sorghum, 

wheat, and other grains until 1984, when its registration as a pesticide was voluntarily canceled. Hexachlorobenzene had also been used in the production of pyrotechnic and ordinance materials for the 

military and in the production of synthetic rubber. 

The general population is not likely to be exposed to large amounts of hexachlorobenzene, but some 

exposure is likely, as many studies have detected small amounts in food and air samples. Traces of hexachlorobenzene have been found in almost all people tested for hexachlorobenzene or its metabolites. 

These amounts of hexachlorobenzene are most likely the result of consumption of low levels in food, with 

a estimated yearly uptake of 68, 22, and 5 μg for adults, toddlers, and infants, respectively. Other sources 

of exposure may include contact with contaminated soil and air. Hexachlorobenzene has been detected in 

agricultural soil at levels up to 440 ng/g, and in lake sediment at #15 ng/g. Ambient air samples have 

usually been reported to range from 0.1 pg/m3

 to 1.5 ng/m3. Hexachlorobenzene has a very low solubility 

in water, so exposure by water is not likely to be significant; ambient water samples have usually been 

below 0.1 parts per trillion (ppt). 

Children are expected to be exposed to hexachlorobenzene by the same routes as adults. Additionally, if 

hexachlorobenzene is present in their mothers, unborn children may be exposed through the placenta and 
HEXACHLOROBENZENE 12 

2. RELEVANCE TO PUBLIC HEALTH 

nursing children may be exposed to hexachlorobenzene present in milk. Human milk samples from the 

general population found hexachlorobenzene concentrations in the range of 11–70 ng/g fat. 

See Chapter 6 for more detailed information regarding concentrations of hexachlorobenzene in 

environmental media. 

2.2 SUMMARY OF HEALTH EFFECTS 

Hexachlorobenzene is a toxic organochlorine that has been shown to cause death, systemic (e.g., liver, 

skin, bone, and thyroid), neurological, developmental, endocrine, and immunological toxicity in humans. 

Animal studies have demonstrated that hexachlorobenzene causes reproductive toxicity and increases the 

risk for cancer formation. The most sensitive target organs for hexachlorobenzene are the liver, ovary, 

and central nervous system. 

A limited number of occupational studies have associated inhalation of hexachlorobenzene with liver 

effects (strongly with increased porphyrins and weakly with hepatocellular carcinoma), immunological 

effects (decreased neutrophil activity, increased immunoglobulins, and susceptibility to infection), and 

renal effects (microproteinuria) Queiroz et al. 1997, 1998a; Richter et al. 1994; Selden et al. 1999). Most 

data for the inhalation effects of hexachlorobenzene in humans were presented by studies of workers from 

an organochlorobenzene factory and the residents of a nearby rural town (Flix, Spain). Exposure to 

hexachlorobenzene (primarily airborne) pollution has been linked with elevated blood levels of hexachlorobenzene and hepatic effects (increased porphyrins and hepatic enzymes), thyroid effects (decreased 

thyroxine levels; weakly with hypothyroidism, goiter, and thyroid cancer), and impaired development of 

locomotor skills in infants. 

Striking epidemiological evidence was found in studies of a population orally exposed to hexachlorobenzene in southeast Anatolia, Turkey. In the 1950s, widespread ingestion of bread made from grain that 

had been treated with hexachlorobenzene as a pesticide caused an epidemic in this region. The ingested 

dose of hexachlorobenzene was estimated to be in the range of 0.05–0.2 g/day, equivalent to 

0.7–2.9 mg/kg/day for an average person. An extremely high (95%) rate of mortality occurred in infants 

under 2 years of age who had been breast fed by mothers who had ingested the contaminated bread. 

Poisoned infants displayed a condition known as pembe yara or "pink sore" because of the associated 

skin lesions (annular erythema). The infant deaths were primarily associated with cardiorespiratory 

failure secondary to this disease. Other clinical symptoms in these infants included weakness and 
HEXACHLOROBENZENE 13 

2. RELEVANCE TO PUBLIC HEALTH 

convulsions. A disease called kara yara or “black sore” was observed most frequently in children 

between the ages of 6 and 15 years, although some younger children and adults were also affected. It 

appeared after approximately 6 months of exposure; symptoms included photosensitivity, skin fragility 

(causing ulcers and scarring), hyperpigmenntation, and hirsutism. There was a 10% mortality rate among 

kara yara patients. These skin lesions were diagnosed as porphyria cutanea tarda, a specific type of 

vesiculobullous porphyria. The porphyrias are a class of inherited and acquired diseases caused by 

enzymatic defects in heme biosynthesis, leading to the generation of porphyrins, which may cause tissue 

damage, especially in the skin. The human studies and supporting animal studies have clearly 

demonstrated that hexachlorobenzene causes porphyria. 

Other symptoms diagnosed in this hexachlorobenzene-exposed population were: loss of appetite, 

weakness, arthritis (a swelling and spindling of the fingers, but with little pain), hepatomegaly, enlarged 

thyroid, and inability to perform simple, everyday activities such as handling eating utensils, rising from a 

squat, and climbing stairs. Clinical findings persisted in most subjects, including high porphyria, dermal 

lesions, multiple neurological effects, skeletomuscular effects, enlarged liver, and enlarged thyroid. In 

adult pregnant women who had been exposed as children, suggestive (but not conclusive) evidence of 

elevated incidences of miscarriages and stillbirths was found. Similar irreversibility has been seen in 

animal studies: developmentally exposed rats exhibited a significantly increased response to startling as 

adults and acutely exposed rats still exhibited porphyria more than 500 days after exposure. 

No studies were located regarding health effects in humans or animals following dermal exposure to 

hexachlorobenzene. However, an acute study in rats suggested that hexachlorobenzene can be absorbed 

across the skin. 

The primary target systems for hexachlorobenzene are hepatic toxicity, reproductive toxicity, 

developmental toxicity, and carcinogenesis; these are discussed below. The reader is referred to 

Section 3.2, Discussion of Health Effects by Route of Exposure, for more detailed information and 

discussions of additional effects. 

Hepatic Effects. The most consistently identified effect following exposure of humans or animals to 

hexachlorobenzene is porphyria. The porphyrias are a class of inherited and acquired diseases caused by 

enzymatic defects in heme biosynthesis, leading to the generation of porphyrins (see Section 3.5, 

Mechanisms of Action). The build-up of high levels of porphyrins in the body is known to cause liver 

(including cirrhosis, siderosis [accumulation of iron], focal necrosis, hyperplasia), and kidney (renal 
HEXACHLOROBENZENE 14 

2. RELEVANCE TO PUBLIC HEALTH 

failure) damage. Additionally, phototoxicity occurs as porphyrins accumulated in the skin are activated 

by sunlight to generate reactive oxygen species, causing tissue damage. As a result, skin lesions occur 

most commonly on areas exposed to sunlight, such as the hands and face. Available data also suggest that 

porphyrins may activate the immune system. Porphyria, diagnosed by the presence of high levels of 

porphyrins in the blood, feces, or urine, has been detected following exposures to hexachlorobenzene in 

workers, in the residents of Flix, Spain (primarily inhalation exposures resulting from a nearby 

organochlorine factory), and in the Turkish epidemic (oral exposure with contaminated grain). Exposed 

patients from the Turkish epidemic also exhibited hepatomegaly. 

Several studies in rats and mice have observed porphyria, but no clear relationship has been established 

between porphyria and other hepatic effects seen in animals, such as peribiliary lymphocytosis and 

fibrosis, hepatomegaly and increased liver weight, enzyme induction, and degenerative pathological 

changes. 

Reproductive Effects. Although no reliable evidence of reproductive toxicity has been observed in 

humans, the reproductive performance of rats has been adversely affected at doses as low as 

16 mg/kg/day of hexachlorobenzene (decreased fertility, increased numbers of stillborn pups). 

Distribution studies have identified the ovaries as a site of hexachlorobenzene accumulation, and 

intermediate-duration exposures to doses of hexachlorobenzene as low 0.01 mg/kg/day have been shown 

to cause ovarian lesions in adult female Cynomolgus monkeys. At higher doses, studies in Cynomolgus 

monkeys, Rhesus monkeys, and rats have reported changes in organ weight, degenerative changes, and 

disruptions in steroidogenesis (estrogen and progesterone). 

Developmental Effects. Human and animal studies have demonstrated that hexachlorobenzene 

crosses the placenta to accumulate in fetal tissues and is transferred in breast milk. 

The poisoning epidemic in Anatolia, Turkey, demonstrated that hexachlorobenzene is a developmental 

toxin. Children exposed under 2 years of age were the most susceptible (95% mortality, skin lesions). 

However, children under 15 were also more susceptible than adults, and exhibited both immediate (10% 

mortality, skin lesions) and persistent (dermal, neurological, skeletomuscular, hepatic, and thyroid effects) 

symptoms. Exposure for adults was estimated at 0.05–0.2 g/day of hexachlorobenzene (in bread made 

from contaminated grain) between 1955 and 1959. 
HEXACHLOROBENZENE 15 

2. RELEVANCE TO PUBLIC HEALTH 

Other human studies investigating developmental toxicity have been limited by small study size and low 

levels of hexachlorobenzene exposure; they have found suggestive evidence of an increased risk of 

undescended testis and impaired development of locomotor skills in newborn babies. 

Animal studies have verified that hexachlorobenzene impaired neurological development and reduced 

neonatal viability and growth. Although an increased risk of undescended testis has not been observed in 

animals, the occurrence of cleft palate, renal agenesis, and minor skeletal abnormalities in mice are 

consistent with a possible teratogenic role for hexachlorobenzene. 

Cancer. Among the general population, case-control studies have generally found no association 

between hexachlorobenzene levels in blood or tissues and incidence of breast or other cancers. Data from 

men exposed to hexachlorobenzene by inhalation (occupationally or as a result of nearby air pollution 

from an organochlorine factory in Flix, Spain) provide weak evidence for an association between 

hexachlorobenzene exposure and cancer of the liver, thyroid, and brain. Because hexachlorobenzene 

produces porphyria, it is noteworthy that several human studies have associated porphyria with the 

increased incidence of liver cancer. 

Several animal studies have demonstrated that oral exposure to hexachlorobenzene increases the 

incidence of tumor formation. The evidence of carcinogenicity is strongest in the liver; 

hexachlorobenzene has been shown to induce hyperplasia (in rats, mice, pigs, dogs, and monkeys), 

metaplasia (in rats), benign tumors (hepatoma in mice and rats; hemagniohepatoma and bile duct 

adenoma in rats), and malignant tumors (hepatocarcinoma in rats, mice, and hamsters; bile duct 

adenocarcinoma in rats). Additionally, exposure to hexachlorobenzene has been shown to induce renal 

metaplasia, adenomas and renal cell carcinomas (in rats, mice, and hamsters); lymphosarcomas (in rats, 

mice, and hamsters); adrenal hyperplasia and pheochromocytoma (in rats); parathyroid adenomas (in 

rats); and hemangioendothelioma and thyroid tumors (in hamsters). In the Ninth Report on Carcinogens, 

the NTP classified hexachlorobenzene as reasonably anticipated to be a human carcinogen. The EPA 

classified hexachlorobenzene as a probable human carcinogen, Group B2, on the basis that oral 

administration of hexachlorobenzene has been shown to induce tumors in the liver, thyroid, and kidney in 

three rodent species. IARC has classified hexachlorobenzene as possibly carcinogenic to humans (Group 

2B), based on inadequate evidence in humans and sufficient evidence in experimental animals for 

carcinogenicity. For more information, see Sections 3.2.1.7 and 3.2.2.7, Cancer. 
HEXACHLOROBENZENE 16 

2. RELEVANCE TO PUBLIC HEALTH 

2.3 MINIMAL RISK LEVELS (MRLs) 

Inhalation MRLs 

Although limited evidence identified organs (liver, kidneys) and systems (endocrine, neurological and 

immune) that may be targets of hexachlorobenzene toxicity following inhalation exposures in humans and 

animals, these data are qualitatively and quantitatively inadequate for use in developing acute, 

intermediate, or chronic inhalation MRLs for hexachlorobenzene. 

Oral MRLs 

C	 An MRL of 0.008 mg/kg/day has been derived for acute-duration oral exposure (14 days or 

less) for hexachlorobenzene. 

This MRL is based on a critical evaluation of a developmental study (Goldey and Taylor 1992) that 

observed a lowest-observed-adverse-effect level (LOAEL) of 2.5 mg/kg/day for hyperactivity in offspring 

rats. An uncertainty factor of 300 was used (10 for extrapolation from rats to humans, 10 for human 

variability, and 3 for use of a minimal LOAEL). 

Human data have shown that the developing central nervous system is a target of hexachlorobenzene 

toxicity. Many breast-fed infants of mothers who ingested hexachlorobenzene-contaminated bread during 

an epidemic of hexachlorobenzene poisoning in Turkey between 1955 and 1959 showed symptoms of 

pembe yara, which included weakness, convulsions, and annular erythema prior to death (Cripps et al. 

1984; Peters et al. 1982, 1987). In children exposed (at an average age of 7 years) to hexachlorobenzenecontaminated grain, neurological effects persisted into adulthood (weakness, paresthesia, sensory shading, 

myotonia, and cogwheeling [irregular jerkiness of movement due to increased muscle tone as seen in 

Parkinson’s disease]) (Cripps et al. 1984; Peters et al. 1982). Additionally, a preliminary report of infants 

from Flix, Spain, found an association between high hexachlorobenzene levels in milk and blood and 

impaired development of locomotor skills (Sala et al. 1999a). 

Toxicology experiments have identified the same targets in developing animals. A developmental study 

in rats (Goldey and Taylor 1992) provided the most appropriate data from which to derive an acuteduration oral MRL for hexachlorobenzene. Virgin female rats were fed 2.5 or 25 mg/kg/day hexachlorobenzene for 4 days, 2 weeks prior to mating with unexposed males, and the developmental neurotoxicity 

of hexachlorobenzene was assessed in offspring using a battery of tests (Goldey and Taylor 1992). Pups 
17 

C

HEXACHLOROBENZENE 

2. RELEVANCE TO PUBLIC HEALTH 

treated with either dose reoriented themselves significantly more quickly in a negative geotaxis test (on 

postnatal days 6 and 8), required less time in an olfactory discrimination test, and demonstrated increased 

exploratory activity in a motor activity test (on postnatal days 9–11). No significant effects on learning 

(swim T-maze) or motor activity (measured in older offspring on postnatal days 40 and 50, respectively) 

were detected. Hexachlorobenzene-exposed offspring at the 25 mg/kg/day dose level exhibited 

significantly altered acoustic startle responses (decreased at 23 days of age and increased at 90 days of 

age compared to controls). Thus, the study identified a LOAEL of 2.5 mg/kg/day for hyperactivity in the 

offspring rats, and this was the most sensitive LOAEL identified for the acute toxicity of 

hexachlorobenzene. 

Other animal studies have demonstrated that exposure to hexachlorobenzene can produce neurological 

effects. In rats exposed prenatally through adulthood, decreased operant learning ability (“post-reinforcement pause” and “index of curvature”) was observed on postnatal day 150 (Lilienthal et al. 1996). 

Moreover, oral exposure to hexachlorobenzene has also been shown to interfere with the function of the 

nervous system, inducing mild reduction in conduction velocity of the sciatic nerve and denervation 

(fibrillations, chronic repetitive discharges) (Sufit et al. 1986), hyperexcitability, tremors, muscle 

fasciculations, clonic convulsions, ataxia, lethargy, and paralysis in adult rats (Kennedy and Wigfield 

1990; Kimbrough and Linder 1974; Koss et al. 1978; Nikolaev et al. 1986; Ockner and Schmid 1961); 

convulsions, tremors, and progressive weakness in litters of female rats (Cripps 1990); tremor in adult 

C57B1/6J mice (Hahn et al. 1988); dysrhythmic electroencephalogram in adult Beagle dogs (Sundlof et 

al. 1981); severe tremors and muscular weakness in adult Rhesus monkeys (Knauf and Hobson 1979); 

hypoactivity, lethargy, and ataxia in infant Rhesus monkeys (Iatropoulos et al. 1978); and tremors, 

panting, and unsteady gait in adult SPF pigs (Den Tonkelaar et al. 1978). 

An MRL of 0.0001 mg/kg/day has been derived for intermediate-duration oral exposure 

(15–364 days) for hexachlorobenzene. 

This MRL is based on a LOAEL of 0.01 mg/kg/day for minimal ovarian effects in monkeys (Babineau et 

al. 1991). An uncertainty factor of 90 was used (3 for extrapolation from monkeys to humans, 10 for 

human variability, and 3 for use of a minimal LOAEL). Ultrastructural studies of ovaries collected from 

monkeys (Babineau et al. 1991; Bourque et al. 1995; Jarrell et al. 1993) provide the most appropriate data 

for use deriving an intermediate oral MRL for hexachlorobenzene. Female Cynomolgus monkeys were 

fed doses of 0.01–10 mg/kg/day of hexachlorobenzene in glucose in gelatin capsules for 90 days; the 

studies focused exclusively on end points relevant to reproductive toxicity (ovarian function and 
HEXACHLOROBENZENE 18 

2. RELEVANCE TO PUBLIC HEALTH 

histopathology). The LOAEL of 0.01 mg/kg/day for reproductive effects is the most sensitive LOAEL 

for the intermediate toxicity of hexachlorobenzene (Babineau et al. 1991). 

Two early studies (Babineau et al. 1991; Jarrell et al. 1993) used doses of 0.1, 1, and 10 mg/kg/day of 

hexachlorobenzene. In all treated animals, hexachlorobenzene decreased the total number of oocytes and 

primary follicles, caused oocyte necrosis and follicular degeneration, and induced histopathological 

changes in the ovarian epithelium (cell stratification; decreased nuclear membrane distinction; increased 

density and granularity of oocyte nuclei; and increased numbers of aggregated lysosomes, vesicles, and 

vacuoles). The severity was dose-dependent. 

The Bourque et al. (1995) follow-up study extended the observation of ultrastructural effects in the ovary 

to 0.01 mg/kg/day. At this dose, mitochondria in developing follicles were condensed and deformed. At 

higher doses (up to 10 mg/kg/day), the mitochondria were progressively more damaged and other changes 

were noted, such as indentation of nuclear membranes and abnormal accumulation of lipid in the 

cytoplasm of follicular cells. Because these effects were not associated with changes in oocyte fertility 

(measured in vitro), they were considered minimally adverse. Thus, these studies identify a LOAEL of 

0.01 mg/kg/day for minimal reproductive effects in the treated monkeys. 

Other monkey studies have also observed evidence of ovarian effects. In female Cynomolgus monkeys 

given capusles with at least 0.1 mg/kg/day of hexachlorobenzene for 90 days, a dose-related decrease in 

serum progesterone levels during the luteal (but not follicular and periovulatory) phase of the menstrual 

cycle, lengthening of the menstrual cycle, ultrastructural changes in surface epithelium of the ovary 

(indicative of cellular degeneration), and changes in ovary surface epithelial cell shape (length to width 

ratio) were detected (Foster et al. 1992a; Sims et al. 1991). In female Rhesus monkeys, gavage doses of 

at least 64 (but not lower doses up to 32) mg/kg/day of hexachlorobenzene for 60 days induced 

degenerative changes of the ovarian follicle, stroma, and germinal epithelium (Iatropoulos et al. 1976), 

and suggestive evidence of unusual steroidogenic activity (depressed serum potassium) was seen in 

Rhesus monkeys given 128 mg/kg/day for at least 60 days (Knauf and Hobson 1979). 

Rat studies have confirmed the reproductive toxicity of hexachlorobenzene in the ovary. Increased serum 

progesterone levels and elevated ovarian weights were observed in superovulated female Sprague-Dawley 

rats orally administered $1 mg/kg/day hexachlorobenzene by gavage for 21 days (Foster et al. 1992b). 
19 

C

HEXACHLOROBENZENE 

2. RELEVANCE TO PUBLIC HEALTH 

Super-ovulated (but not normal cycling) female Sprague-Dawley rats gavaged with 50 mg/kg/day of 

hexachlorobenzene for 5 days exhibited significant elevation of serum levels of progesterone (Foster et al. 

1993). In a subsequent study with ovariectomized female Sprague-Dawley rats, gavage doses of at least 

1 mg/kg/day for 30 days significantly decreased circulating corticosterone and cortisol levels, without 

affecting levels of circulating aldosterone and progesterone levels or adrenal gland weight (Foster et al. 

1995a). The investigators concluded that hexachlorobenzene exposure induces alterations in 

steroidogenesis of cells of the adrenal cortex inner zone. 

Although clear evidence of reproductive toxicity has not been observed in humans, suggestive data 

indicated that hexachlorobenzene may increase the risk for spontaneous abortion (miscarriages and 

stillbirths). Therefore, the intermediate-duration oral MRL is considered relevant to human health. 

An MRL of 0.00005 mg/kg/day has been derived for chronic-duration oral exposure (365 days 

or more) for hexachlorobenzene. 

This MRL was based on a critical evaluation of a multigenerational study (Arnold et al. 1985), which 

observed a LOAEL of 0.016 mg/kg/day for hepatic effects in F1 male rats. An uncertainty factor of 

300 was used (10 for extrapolation from rats to humans, 10 for human variability, and 3 for use of a 

minimal LOAEL). 

Although studies designed to detect liver histopathology have not been conducted in humans, 

epidemiological studies have detected hepatic effects as indicated by increased porphyrin and enzyme 

levels in people exposed for chronic durations to hexachlorobenzene (Herrero et al. 1999; Sala et al. 

1999b; Seldon et al. 1999). Analysis of the available human and animal chronic oral toxicity data for 

hexachlorobenzene indicated that a 130-week study in rats reported by Arnold et al. (1985) provided the 

most appropriate data for use in the development of an oral chronic MRL for hexachlorobenzene. 

Briefly, male and female weanling Sprague-Dawley rats were fed 0, 0.016, 0.08, 0.4, or 2 mg/kg/day of 

hexachlorobenzene for 3 months prior to mating and through weaning of F1 offspring, when they were 

sacrificed. The F1 offspring were continued on their parents’ diet from weaning throughout the remainder 

of their lives (130 weeks). Statistically significant increases in the incidences of peribiliary 

lymphocytosis and fibrosis were observed in the livers of male rats at $0.016 mg/kg/day in the F1 

generation lifetime study. Significant dose-dependent trends were also found for hepatic basophilic 

chromogenesis at $0.4 mg/kg/day in both genders of rats. The 2 mg/kg/day dose group rats also 

exhibited increased pup mortality (in both genders) and severe chronic nephrosis (in males only). 
HEXACHLOROBENZENE 20 

2. RELEVANCE TO PUBLIC HEALTH 

The increased incidences of peribiliary lymphocytosis and fibrosis in treated males were considered to 

represent a minimal effect. These are common spontaneous lesions in aging rats and occurred in 

approximately 30% of controls in this study. For peribiliary fibrosis, incidence was increased in all 

treated groups (statistically significant in the 0.016 and 2 mg/kg/day groups), but there was no clear 

evidence of a dose-response (13/48, 23/48, 21/48, 21/49, and 23/49, respectively, in the control, 0.016, 

0.08, 0.4, and 2 mg/kg/day groups). For peribiliary lymphocytosis, the incidence was increased in all 

treated groups (statistically significant in the 0.016, 0.08, and 2 mg/kg/day groups), and the trend was also 

statistically significant (16/48, 27/48, 26/48, 21/49, and 32/49, respectively, in the control, 0.016, 0.08, 

0.4, and 2 mg/kg/day groups). Because incidences of these lesions in the control and treated females were 

similar to the control males (ranging from 6/49 to 14/49), the incidence levels in control males do not 

appear unusually low. Overall, these findings suggest that hexachlorobenzene produced a minimal 

hepatic effect in male rats at the lowest doses administered by increasing the incidence of age-related 

hepatic lesions. The LOAEL of 0.016 mg/kg/day reported in this study for hepatic effects is the most 

sensitive LOAEL for the characteristic chronic toxicity of hexachlorobenzene, liver toxicity. However, 

the study did not identify a no-observed-adverse-effect level (NOAEL). 

Numerous animal studies have clearly demonstrated that the liver is a major target organ of hexachlorobenzene exposure. Short-term experiments in rats have observed increased liver weight, increased hepatic 

porphyrins, liver histopathology (cytoplasmic vacuolation, anisokaryosis, and pyknotic hepatocytes), and 

increased serum cholesterol (Krishnan et al. 1991; Lecavalier et al. 1994; Rajamanickam and Padmanaban 

1974; Richter et al. 1981). Intermediate- and chronic-duration experiments in rats have observed these 

and other signs, including additional liver histopathology (degeneration, hypertrophic hepatocytes with 

eosinophilic cytoplasm with thready basophilic structures, as well as inflammatory cell infiltrates), 

decreased liver retinoid levels, and elevated liver enzymes (Andrews et al. 1989, 1990; Den Besten et al. 

1993; Elder and Urquhart 1986; Kennedy and Wigfield 1990; Koss et al. 1978, 1983; Kuiper-Goodman et 

al. 1977; Ockner and Schmid 1961; Smith and Cabral 1980; van Raaij et al. 1993b). Similar effects have 

been observed in intermediate-duration experiments in Rhesus monkeys (Iatropoulos et al. 1976; Knauf 

and Hobson 1979), beagle dogs (Sundlof et al. 1981), and pigs (Den Tonkelaar et al. 1978). 

HEXACHLOROBENZENE 22 
HEXACHLOROBENZENE 23 

3. HEALTH EFFECTS 

3. HEALTH EFFECTS 

3.1 INTRODUCTION 

The primary purpose of this chapter is to provide public health officials, physicians, toxicologists, and 

other interested individuals and groups with an overall perspective on the toxicology of hexachlorobenzene. It contains descriptions and evaluations of toxicological studies and epidemiological 

investigations and provides conclusions, where possible, on the relevance of toxicity and toxicokinetic 

data to public health. 

A glossary and list of acronyms, abbreviations, and symbols can be found at the end of this profile. 

3.2 DISCUSSION OF HEALTH EFFECTS BY ROUTE OF EXPOSURE 

To help public health professionals and others address the needs of persons living or working near 

hazardous waste sites, the information in this section is organized first by route of exposure (inhalation, 

oral, and dermal) and then by health effect (death, systemic, immunological, neurological, reproductive, 

developmental, genotoxic, and carcinogenic effects). These data are discussed in terms of three exposure 

periods: acute (14 days or less), intermediate (15–364 days), and chronic (365 days or more). 

Levels of significant exposure for each route and duration are presented in tables and illustrated in 

figures. The points in the figures showing no-observed-adverse-effect levels (NOAELs) or 

lowest-observed-adverse-effect levels (LOAELs) reflect the actual doses (levels of exposure) used in the 

studies. LOAELs have been classified into "less serious" or "serious" effects. "Serious" effects are those 

that evoke failure in a biological system and can lead to morbidity or mortality (e.g., acute respiratory 

distress or death). "Less serious" effects are those that are not expected to cause significant dysfunction 

or death, or those whose significance to the organism is not entirely clear. ATSDR acknowledges that a 

considerable amount of judgment may be required in establishing whether an end point should be 

classified as a NOAEL, "less serious" LOAEL, or "serious" LOAEL, and that in some cases, there will be 

insufficient data to decide whether the effect is indicative of significant dysfunction. However, the 

Agency has established guidelines and policies that are used to classify these end points. ATSDR 

believes that there is sufficient merit in this approach to warrant an attempt at distinguishing between 

"less serious" and "serious" effects. The distinction between "less serious" effects and "serious" effects is 

considered to be important because it helps the users of the profiles to identify levels of exposure at which 
HEXACHLOROBENZENE 24 

3. HEALTH EFFECTS 

major health effects start to appear. LOAELs or NOAELs should also help in determining whether or not 

the effects vary with dose and/or duration, and place into perspective the possible significance of these 

effects to human health. 

The significance of the exposure levels shown in the Levels of Significant Exposure (LSE) tables and 

figures may differ depending on the user's perspective. Public health officials and others concerned with 

appropriate actions to take at hazardous waste sites may want information on levels of exposure 

associated with more subtle effects in humans or animals (LOAELs) or exposure levels below which no 

adverse effects (NOAELs) have been observed. Estimates of levels posing minimal risk to humans 

(Minimal Risk Levels or MRLs) may be of interest to health professionals and citizens alike. 

Levels of exposure associated with carcinogenic effects (Cancer Effect Levels, CELs) of hexachlorobenzene are indicated in Table 3-2 and Figure 3-2. Because cancer effects could occur at lower exposure 

levels, Figure 3-2 also shows a range for the upper bound of estimated excess risks, ranging from a risk of 

1 in 10,000 to 1 in 10,000,000 (10-4 to 10-7), as developed by EPA. 

Estimates of exposure levels posing minimal risk to humans (Minimal Risk Levels or MRLs) have been 

made for hexachlorobenzene. An MRL is defined as an estimate of daily human exposure to a substance 

that is likely to be without an appreciable risk of adverse effects (noncarcinogenic) over a specified 

duration of exposure. MRLs are derived when reliable and sufficient data exist to identify the target 

organ(s) of effect or the most sensitive health effect(s) for a specific duration within a given route of 

exposure. MRLs are based on noncancerous health effects only and do not consider carcinogenic effects. 

MRLs can be derived for acute, intermediate, and chronic duration exposures for inhalation and oral 

routes. Appropriate methodology does not exist to develop MRLs for dermal exposure. 

Although methods have been established to derive these levels (Barnes and Dourson 1988; EPA 1990), 

uncertainties are associated with these techniques. Furthermore, ATSDR acknowledges additional 

uncertainties inherent in the application of the procedures to derive less than lifetime MRLs. As an 

example, acute inhalation MRLs may not be protective for health effects that are delayed in development 

or are acquired following repeated acute insults, such as hypersensitivity reactions, asthma, or chronic 

bronchitis. As these kinds of health effects data become available and methods to assess levels of 

significant human exposure improve, these MRLs will be revised. 
HEXACHLOROBENZENE 25 

3. HEALTH EFFECTS 

A User's Guide has been provided at the end of this profile (see Appendix B). This guide should aid in 

the interpretation of the tables and figures for Levels of Significant Exposure and the MRLs. 

3.2.1 Inhalation Exposure 

3.2.1.1 Death 

No studies were located regarding death in humans or animals following inhalation exposure to hexachlorobenzene. 

3.2.1.2 Systemic Effects 

No studies were located regarding respiratory, cardiovascular, gastrointestinal, hematological, 

musculoskeletal, dermal, or ocular effects in humans or animals following inhalation exposure to hexachlorobenzene. 

Hepatic Effects. It is well known that ingestion of hexachlorobenzene can produce porphyria (see 

Section 3.2.2.2), but there are few data to suggest a similar effect from inhalation exposure. There was a 

small, but statistically significant, increase (in comparison to matched controls) in total urinary porphyrins 

in a group of nine workers exposed for 1–19 years to hexachlorobenzene and other organochlorine 

compounds (including octachlorostyrene, another porphyrinogenic substance [Smith et al. 1986d]) while 

using hexachloroethane as an aluminum degassing agent at aluminum smelters (Selden et al. 1999). The 

prevalence of porphyria-related symptoms did not differ between groups. Serum levels of hexachlorobenzene were positively correlated to urine porphyrin levels, but urine porphyrin levels also correlated, to 

a similar degree, with serum octachlorostyrene levels. Therefore, this study cannot ascertain whether the 

preclinical porphyria observed in aluminum smelter workers is due to hexachlorobenzene, 

octachlorostyrene, or both chemicals. End points of hepatic toxicity have been investigated in residents 

of Flix, Spain, a small village in Catalonia; a nearby electrochemical factory that manufactures 

organochlorines has been implicated in the village’s high environmental levels of hexachlorobenzene 

(35 ng/m3

 as a 24-hour average in air). Analysis of blood hexachlorobenzene and urinary porphyrins in 

604 residents of Flix, including 185 factory workers, showed that blood hexachlorobenzene levels were 

roughly 5-fold higher in factory workers than in other residents. However, there were no cases of clinical 

porphyria cutanea tarda in either group and no evidence that preclinical porphyria was more prevalent in 

the factory workers than in other residents or that there was any association between urinary porphyrin 
HEXACHLOROBENZENE 26 

3. HEALTH EFFECTS 

levels and blood hexachlorobenzene levels (Herrero et al. 1999; Sala et al. 1999b). A subsequent study 

verified the 5-fold increase in 75 factory workers, and detected a positive, statistically significant 

correlation between blood hexachlorobenzene levels and serum γ-glutamyltransferase (Sala et al. 2001a). 

However, no correlation was seen for serum levels of either aspartate or alanine aminotransferase. The 

NOAEL for liver effects in the village residents are shown in Table 3-1 and Figure 3-1. In 52 workers 

employed for 1–25 years at a chemical plant where hexachlorobenzene was the primary byproduct, 

compared with controls drawn from a local blood bank (Queiroz et al. 1998a), serum aspartate and 

alanine transaminase levels were increased but no correlation was detected between serum levels of 

transaminase and blood hexachlorobenzene. The available human data, while suggestive, have not 

conclusively shown hepatic effects due to inhaled hexachlorobenzene. No studies were located regarding 

the hepatic effects of airborne hexachlorobenzene in animals. 

Renal Effects. Based on analysis of blood samples from 608 individuals, workers (n=189) at an 

electrochemical factory near Flix, Spain did not have a significantly increased risk of high serum 

creatinine, a marker for glomerular disease, in comparison to other Flix residents (n=419) despite hexachlorobenzene blood levels that were approximately 5-fold higher (Sala et al. 1999b). However, the 

power of this study to find an effect was limited by the small sample size (only nine individuals with high 

serum creatinine were found in the study population). “Marked changes in kidney functions” including 

microproteinuria were observed in Czechoslovakian workers with high blood hexachlorobenzene levels 

following occupational inhalation exposure to hexachlorobenzene, originally at 2.1–10.8 mg/m3

 and then 

at 0.012–0.022 mg/m3

, from 1983 to 1990 (Richter et al. 1994). No studies were located regarding renal 

effects of inhaled hexachlorobenzene in animals. 

Endocrine Effects. Although oral exposure to hexachlorobenzene has been clearly associated with 

thyroid effects, only limited information is available regarding the endocrine effects of hexachlorobenzene following inhalation exposure. A study of residents (192–558, depending on end point) of Flix, 

Spain, where a nearby electrochemical factory had resulted in high air and blood levels of hexachlorobenzene, detected statistically significant correlations between increased blood hexachlorobenzene levels 

and decreased levels of total thyroxine (T4); serum levels of thyroid stimulating hormone (TSH) were not 

affected (Sala et al. 2001a). Similarly, an earlier study did not detect changes in serum levels of TSH in 

189 workers at the factory in comparison to 419 other Flix residents despite 5-fold higher blood hexachlorobenzene levels in the workers (Sala et al. 1999b). Both studies were limited by small sample size, 
Table 3-1. Levels of Significant Exposure to Hexachlorobenzene - Inhalation I 

Key toa 

figure 

Species 

(strain) 

Exposure1 

duration/ 

frequency System 

NOAEL 

(mglm3) 

Less serious 

(mg/m3) 

LOAEL 

Serious 

(mg/m3) 

Reference 

Chemical Form 

IJ 

z 

;21 

1 

ACUTE EXPOSURE 

lmmunologicallLymphoreticular 

Rat 1-4 d 

(Sprague- 4 hr/d 

Dawley) 

4.4 M 33 M (slight impairment of 

pulmonary immune 

defenses) 

Sherwood et al. 1989 

rn 

rn 

z 

w 

I 

rn

5

I 

n

n 

0 

rn 

rn 

d 

is 
Table 3-1. Levels of Significant Exposure to Hexachlorobenzene - Inhalation (continued) I 

p!

E 

Exposure1 LOAEL 

Key toa Species duration/ NOAEL Less serious Serious Reference 

figure (strain) frequency System (mg/m3) (mglm3) (mglm3) Chemical Form 

INTERMEDIATE EXPOSURE 

lmmunological/Lymphoreticular 

2 Rat 4 wk 35 M (slight impairment of Sherwood et al. 1989 

(SpragueDawley) 4 dwk 

4 hr/d 

pulmonary immune 

defense) 

w 

I 

rn

Fi

I 

N 

CD 
Table 3-1. Levels of Significant Exposure to Hexachlorobenzene - Inhalation (continued) 

Exposurd LOAEL 

Key to" Species duration/ NOAEL Less serious Serious Reference 

figure (strain) frequency System (mglm3) (mglm3) (mglm3) Chemical Form 

CHRONIC EXPOSURE 

Systemic 

3 Human Herrero et al. 1999 40yr Hepatic 0.000035 

~~ ~ 

'The number corresponds to entries in Figure 3-1. 

d = day(s); hr = hours(s); LOAEL = lowest-observable-adverse-effect level; M = male; NOAEL = no-observable-adverse-effect level; wk = week(s); yr = year(s) 

w 

I 

F

5

I 

rn 

n 

n 

rn 

9

v) 
Figure 3-1. Levels of Significant Exposure to Hexachlorobenzene - Inhalation 

Acute (114 days) 

mg/rn3 

100 

w 

I 

F

5

I 

c-Cat -Humans f-Ferret n-Mink Cancer Effect Level-Animals Cancer Effect Level-Humans LD50lLC50 

d-Dog k-Monkey j-Pigeon o-Other LOAEL, More Serious-Animals A LOAEL, More Serious-Humans I Minimal Risk Level W

r-Rat rn-Mouse e-Gerbil (3LOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans I foreffects 

p-Pig h-Rabbit s-Hamster 0NOAEL -Animals NOAEL- Humans other than 

q-Cow a-Sheep g-Guinea Pig Cancer 

I 

0 
100 

Figure 3-1. Levels of Significant Exposure to Hexachlorobenzene - Inhalation (Continued) 

rnglrn3 

Q2r 

1C 

c-Cat -Humans 

d-Dog k-Monkey r-Rat m-Mouse 

p-Pig h-Rabbit 

._9:!9!!-a-Sl?_eeP_ -

Intermediate (15-364 days) 

f-Ferret n-Mink Cancer Effect Level-Animals 

j-Pigeon o-Other 

e-Gerbil 

s-Hamster . 

0 LOAEL, More Serious-Animals 

(3LOAEL, Less Serious-Animals 

0NOAEL -Animals 

-9:GULneaP'g .. . 

Cancer Effect Level-Humans 

A LOAEL, More Serious-Humans 

A LOAEL, Less Serious-Humans 

A NOAEL - Humans 

- __ ._.___ - __ --

LD50/LC50

I ;Minimal Risk for effects Level A other than 

Cancer .__-----

u 

I 

rn

F 

-I

I 

m 

n 

n 

rn 

0 

-i 

cn 

I

F

0

E 

7J 

rn 

z 

N 

rn 

z 

rn 
0.0001 

Figure 3-1. Levels of Significant Exposure to Hexachlorobenzene - Inhalation (Continued) 

Chronic (2365days) 

Systemic 

rnglrn3 

A3 

1E-5 

c-Cat -Humans f-Ferret n-Mink Cancer Effect Level-Animals 

d-Dog

r-Rat 

p-Pig

q-Cow 

k-Monkey

m-Mouse 

h-Rabbit 

a-Sheep 

jPigeon o-Other 

e-Gerbil 

s-Hamster 

g-Guinea Pig 

LOAEL, More Serious-Animals 

LOAEL, Less Serious-Animals 

0NOAEL -Animals 

Cancer Effect Level-Humans 

A LOAEL, More Serious-Humans 

A LOAEL, Less Serious-Humans 

A NOAEL- Humans 

I

: 

0

F

0 

;o 

rn

z

N 

rn

Z 

rn 

w 

I 

rn

5

I 

rn 

n 

n 

m 

LD50/LC50

: Minimal Risk Level W

: for effects N 

A other than 

Cancer 
HEXACHLOROBENZENE 33 

3. HEALTH EFFECTS 

as only 10 cases of elevated TSH were observed in each of the study populations. A larger survey found 

that the prevalences of goiter and hypothyroidism were similar in ever exposed workers (n=507) and 

nonexposed residents (n=1,293) (Sala et al. 1999b). No studies were located regarding endocrine effects 

in animals following inhalation exposure to hexachlorobenzene. 

3.2.1.3 Immunological and Lymphoreticular Effects 

Occupational studies indicate that inhaled hexachlorobenzene may cause physiological changes in 

immune parameters, but these effects are not clearly toxic. Queiroz et al. (1997, 1998a, 1998b) studied a 

group of Brazilian workers (n=51–66) employed for up to 25 years at a chemical plant whose primary 

waste product was hexachlorobenzene. Although it is not clear from the report, workers were presumably 

exposed to hexachlorobenzene by inhalation. Findings in blood samples of exposed workers compared to 

controls selected from a local blood bank were significant decreases in neutrophil chemotaxis, impaired 

neutrophil cytolytic activity (but not phagocytic activity) and significant increases in serum 

immunoglobulins (IgG and IgM, but not IgA). Blood hexachlorobenzene levels were elevated in exposed 

workers, but were not correlated with changes in immunological parameters. Serum immunoglobulins 

(IgG, IgA, IgM, and IgE, but not IgD) were also increased in Czechoslovakian workers with high blood 

hexachlorobenzene levels who had been exposed to hexachlorobenzene in the workplace air from 1983 to 

1990, originally at 2.1–10.8 mg/m3

 and then at 0.012–0.022 mg/m3 (Richter et al. 1994). Immunological 

parameters and serum organochlorine levels were measured in a group of 141 German medical patients 

presenting with variety of acute symptoms (mainly lack of concentration, exhaustion, and common cold) 

who had been occupationally exposed (as teachers, construction workers, and telecommunication 

technicians) for at least 6 months to multiple organochlorines (Daniel et al. 2001). A strong, statistically 

significant, association was detected between high blood levels of hexachlorobenzene and decreased 

levels of interferon-γ blood (IFN-γ) levels. Moreover, patients with low overall organochlorine levels had 

elevated IFN-γ levels. The authors note that IFN-γ is important for increasing the secretion of 

immunoglobulins by plasma cells, and speculate that decreased IFN-γ might increase susceptibility to 

infection. It is not clear whether hexachlorobenzene exposure is responsible for the observed effects, 

because significant cross-correlations were detected between hexachlorobenzene levels and several 

polycyclic biphenyl compounds. 

The animal data provide weak support for an immunological effect of inhaled hexachlorobenzene. 

Observations in male rats exposed to 33–35 mg/m3

 of hexachlorobenzene aerosol for durations ranging 

from 4 hours to 4 weeks (4 days/week, 4 hours/day) included a slight decrease in pulmonary macrophage 
HEXACHLOROBENZENE 34 

3. HEALTH EFFECTS 

bactericidal activity to inhaled Klebesiae pneumoniae, a slight increase in phagocytic activity of alveolar 

(but not peritoneal) macrophages in vitro, and altered lymphocyte mitogenesis induced by T-cell 

(phytohemagglutinin [PHA]) and B-cell (Salmonella typhimurium lipopolysaccharide) mitogens in lungassociated and mesenteric lymph nodes in vitro (Sherwood et al. 1989). No effects were found at 

4.4 mg/m3

 (1-day exposure only). The authors concluded that exposure to hexachlorobenzene at about 

35 mg/m3

 resulted in slight changes to humoral and pulmonary cellular defenses. However, the reported 

changes were only marginally different from controls, the magnitude of the reported effects did not 

generally increase with exposure duration, and some of the results were contradictory (e.g., there was a 

significant increase in PHA-induced mitogenesis in lung-associated lymph nodes, but a significant 

decrease in PHA-induced mitogenesis in mesenteric lymph nodes, in rats exposed to hexachlorobenzene 

for 4 weeks). The NOAEL and LOAEL values from this experiment are recorded in Table 3-1 and 

plotted in Figure 3-1. 

3.2.1.4 Neurological Effects 

No clear evidence of neurological effects following inhalation exposure to hexachlorobenzene is 

available. The prevalence of Parkinson’s disease was not significantly increased in workers at an 

electrochemical factory near Flix, Spain (4/507) compared with other Flix residents (4/1,293) despite 

5-fold higher blood hexachlorobenzene levels in the workers (Sala et al. 1999b). The small number of 

cases is a limitation of this study. A preliminary report (Sala et al. 1999a), based on 63 cases, detected a 

statistically significant association between prenatal hexachlorobenzene exposure and impaired 

development of locomotor skills in newborn babies in Flix, compared with those of nearby villages. A 

different study conducted in Oswego, New York was unable to correlate hexachlorobenzene levels in 

umbilical blood or breast milk with infant intelligence test results (Darvill et al. 2000). The apparent 

discrepancy between these studies may be explained by differences in exposure. No studies were located 

regarding neurological effects in animals following inhalation exposure to hexachlorobenzene. 

3.2.1.5 Reproductive Effects 

No studies were located regarding reproductive effects in humans or animals following inhalation 

exposure to hexachlorobenzene. 
HEXACHLOROBENZENE 35 

3. HEALTH EFFECTS 

3.2.1.6 Developmental Effects 

Developmental effects (spontaneous abortions, low birth weight, and congenital malformations) occurred 

with a similar prevalence among females who ever worked at an electrochemical factory near Flix, Spain 

(n=46–60 for the different end points) and never exposed female residents (n=719–936), despite 5-fold 

higher blood hexachlorobenzene levels in the workers (Sala et al. 1999b). The small number of women 

factory workers is a limitation of this study. No studies were located regarding developmental effects in 

animals by inhalation exposure to hexachlorobenzene. 

3.2.1.7 Cancer 

In comparison to the surrounding Province of Tarragona, the incidences of thyroid cancer and soft-tissue 

sarcoma were significantly increased, and brain tumors marginally increased, for the years 1980–1989 in 

male residents of Flix, Spain, where a nearby organochlorine factory had produced high levels of hexachlorobenzene in the ambient air for decades (40 measurements in 1989–1992 averaged 35 ng/m3

, and the 

researchers suspected concentrations had been higher in years past) (Grimalt et al. 1994). Tumor 

incidences were not elevated in female Flix residents, but exposures of the females (few of whom worked 

at the factory) may have been considerably lower than those of the males (many of whom, including all 

those with tumors, worked at the factory). The findings in males were based on very small numbers of 

observed cases (2–4 for the various tumor types), and were not duplicated in a companion analysis of 

cancer mortality reported in the same paper. Therefore, this study was not conclusive. Hepatocellular 

carcinoma was diagnosed in 1985 in a 65-year-old male who had been exposed to airborne hexachlorobenzene and, to a lesser extent, other organochlorine compounds (e.g., chlorinated benzenes, 

chlorophenols, dioxins, and dibenzofurans) at an aluminum smelter from 1967 to 1973 while using 

hexachloroethane as an aluminum degassing agent (Selden et al. 1989). This finding is suggestive, but 

does not provide rigorous evidence for an association between tumor development and inhalation 

exposure to hexachlorobenzene. No other data were located specifically associating cancer with exposure 

to hexachlorobenzene, but several studies have investigated an apparent association between porphyria 

and subsequent development of liver cancer in humans. These data are relevant because hexachlorobenzene is porphyrogenic in humans. However, factors such as cancer evolution times, liver pathology 

(hepatitis viral infection, fibrosis, or cirrhosis), and age were found to be better predictors of subsequent 

tumor development than porphyrin status in these studies (Axelson 1986; Keczkes and Barker 1976; 

Salata et al. 1985; Siersema et al. 1992; Topi et al. 1980; Waddington 1972). 
HEXACHLOROBENZENE 36 

3. HEALTH EFFECTS 

Animal data regarding the carcinogenicity of hexachlorobenzene via the inhalation route were not 

located, although there is evidence that hexachlorobenzene produces liver tumors in animals by the oral 

route (see Section 3.2.2.8). 

3.2.2 Oral Exposure 

3.2.2.1 Death 

Evidence of human lethality following oral exposure to hexachlorobenzene is derived mainly from 

epidemiologic data from from 1955 to 1959. An estimated 3,000–4,000 people ingested bread prepared 

from grain treated with fungicides composed of 10% hexachlorobenzene, at an estimated 2 kg/1,000 kg 

wheat. There was an extremely high rate of mortality in breast fed children (under 2 years of age) of 

mothers known to have ingested this bread. All children born to porphyric mothers during that epidemic 

died (Gocmen et al. 1989; Peters et al. 1982) and an estimated 1,000–2,000 infants died due to a condition 

known as pembe yara or "pink sore" because of the associated skin lesions (blistering and epidermolysis 

and annular erythema) (Cripps et al. 1984; Peters et al. 1982, 1987). Although a 10% rate of mortality in 

exposed adults has been reported, it is not clear how that figure relates to the expected mortality rates for 

comparable cohorts (Peters et al. 1982, 1987). No quantitation of exposure (dose and duration) was 

presented in any of these clinical reports. However, an estimated dose of 0.05–0.2 g/day 

(0.7–2.9 mg/kg/day for a 70-kg person) is considered to be reliable by the original investigators of the 

Turkey epidemic (Cam and Nigogosyan 1963). Analysis of human milk from exposed women and 

unexposed controls in this epidemic showed hexachlorobenzene concentrations of 0.51 and 0.07 ppm, 

respectively (Cripps et al. 1984; Gocmen et al. 1989; Peters et al. 1982, 1987). A report on an 

epidemiological study in New South Wales, Australia, which gave hexachlorobenzene concentrations 

ranging from trace amounts to 8.2 ppm in human body fat and #0.41 ppb in whole blood, found no 

adverse health effects or mortality associated with these levels of body burden (Brady and Siyali 1972; 

Siyali 1972). 

The acute lethality of ingested hexachlorobenzene in animal studies is relatively low. LD50 data are 

limited to a report from the Russian literature of single-dose oral LD50 values of 3,500 mg/kg for rats, 

4,000 mg/kg for mice, 2,600 mg/kg for rabbits, and 1,700 mg/kg for cats (Savitskii 1964, 1965). 

Although these LD50 values were determined under unknown experimental conditions, they are consistent 

with the remainder of the database showing low acute lethality of hexachlorobenzene. One death was 

observed among 10 rats given a single dose of 600 mg/kg by gavage in corn oil (Lecavalier et al. 1994), 
HEXACHLOROBENZENE 37 

3. HEALTH EFFECTS 

but it is not clear from the report that the death was due to hexachlorobenzene. Lethal levels in animal 

studies are progressively lower as exposure duration is increased. Lethal levels ranged from 19 to 

245 mg/kg/day in most (the exception is discussed below) intermediate-duration feeding studies (Cuomo 

et al. 1991; De Matteis et al. 1961; Den Besten et al. 1993; Den Tonkelaar et al. 1978; Kennedy and 

Wigfield 1990; Kimbrough and Linder 1974; Kuiper-Goodman et al. 1977; Ockner and Schmid 1961) 

and from 6 to 16 mg/kg/day in chronic oral studies (Cabral et al. 1977, 1979; Gralla et al. 1977a). Death 

in these studies was closely associated with occurrence of neurological symptoms (e.g., tremors, paresis, 

weakness, convulsions) and, in some cases, weight loss (Cabral et al. 1977; Cuomo et al. 1991; De 

Matteis et al. 1961; Den Besten et al. 1993; Den Tonkelaar et al. 1978; Gralla et al. 1977a; Ockner and 

Schmid 1961). 

Aside from duration of exposure, other factors that appear to influence susceptibility to hexachlorobenzene-induced mortality include species, strain, sex, age, pregnancy status, diet, nutritional status 

(fasting verus normal diet), and dosing protocol (including the vehicle used and the method of exposure). 

A direct comparison of multiple species was performed by De Matteis et al. (1961), who treated rats, 

mice, guinea pigs, and rabbits with 5,000 ppm of hexachlorobenzene in the diet, providing estimated 

doses of 250, 650, 200, and 245 mg/kg/day, respectively. Guinea pigs, with the lowest estimated daily 

dose, and mice, with the highest dose, were the most severely affected of the species tested, with severe 

neurological effects and death occurring as soon as 8–10 days after the start of exposure. Rats and rabbits 

also developed neurological symptoms and died, but only after $8 weeks of exposure. The severe effects 

in guinea pigs despite the low dose suggest that this species may be especially sensitive to hexachlorobenzene; however, additional supporting data are lacking. Although death was reported in both male and 

female animals in various studies, studies that included both sexes generally reported a higher incidence 

of mortality in females than in males treated with the same doses (Gralla et al. 1977a; Kimbrough and 

Linder 1974; Kuiper-Goodman et al. 1977). Pregnant females in particular seem to be especially 

susceptible to hexachlorobenzene-induced mortality. Two pregnant rats fed hexachlorobenzene during 

gestation were much more severely affected than the nonpregnant rats in the De Matteis et al. (1961) 

study, with one dying before giving birth and the other dying 4 weeks after giving birth. Grant et al. 

(1977) observed death of pregnant dams at doses as low as 8 mg/kg/day in an intermediate-duration 

reproduction study, which is considerably lower than the lethal dose range to nonpregnant animals in 

other intermediate-duration studies (19–245 mg/kg/day). 

Toxicity of hexachlorobenzene is enhanced by use of an oil vehicle in animal studies. This was 

demonstrated by Kennedy and Wigfield (1990), who fed female Wistar rats a diet to which hexachloro­
HEXACHLOROBENZENE 38 

3. HEALTH EFFECTS 

benzene was added either in corn oil or as crystalline chemical. At 1,000 ppm (50 mg/kg/day), 2 deaths 

occurred within 23 days of the start of exposure in the corn oil group and the remaining 27 animals were 

removed from the study 2 days later due to obvious ill health. No deaths occurred in the crystalline 

chemical group (n=27) fed the same dose for 56 days. The increased toxicity of hexachlorobenzene fed 

with corn oil appeared to be related to increased accumulation of the chemical in the body (measured in 

liver, kidney, and spleen) under these conditions. 

Reliable LOAEL values for mortality in each species and duration category are presented in Table 3-2 

and plotted in Figure 3-2. 

3.2.2.2 Systemic Effects 

The available data in humans and laboratory animals indicate that the liver, and specifically, the heme 

biosynthesis pathway, is the major systemic target of hexachlorobenzene toxicity. Human data have also 

shown effects on other systemic targets, including the skin, bone, and thyroid. These effects were less 

common than inhibition of heme biosynthesis in exposed people. No information was located regarding 

respiratory, cardiovascular, gastrointestinal, hematological, renal, ocular, or body weight effects in 

humans following oral exposure to hexachlorobenzene. However, animal data are available for these 

systemic effects and suggest that the blood, lungs, and kidneys may be additional systemic targets of 

hexachlorobenzene. 

The highest NOAEL values and all reliable LOAEL values for systemic effects in each species and 

duration category are presented in Table 3-2 and plotted in Figure 3-2. 

Respiratory Effects. No studies were located regarding respiratory effects of oral hexachlorobenzene exposure in humans. 

Animal studies have shown that ingested hexachlorobenzene can produce pathological effects in the 

lungs. The most widely reported lesions were hypertrophy and proliferation of the lining endothelial cells 

of the pulmonary venules and intra-alveolar accumulation of foamy-looking macrophages. The foamy 

appearance of macrophages is a result of increased lipid content (Goldstein et al. 1978). These 
\ 

I Table 3-2. Levels of Significant Exposure to Hexachlorobentene - Oral 

Exposurd LOAEL 0

I 

duration/ 5 

~ey toa Species frequency NOAEL Less serious Serious Reference a3

0 figure (Strain) (Specific route) System (mglkglday) (mglkglday) (mglk@day) Chemical Form W 

rn 

z 

N ACUTE EXPOSURE rn

z 

Death rn 

650 F (death) De Matteis et at. 1961 

200 F (death) De Matteis et al. 1961 

Systemic 

Billi de Catabbi et al. 3 Rat 1,2,3, or 4 Hepatic 1000 F (increased porphyria, 2000a I altered hepatic enzyme (wistar) wks F 5 d/wk levels) 5

I 

rn 

-n 

4 Rat 1 or 7 wks, Hepatic 1000 F (increased urinary Billi de Catabbi et al. 

0 

4 

2000a (Wistar) # of doses per porphyrin excretion) 2

wk unknown ' 

5 Rat 2wk Renal 100 M (incr kidney wt, Bouthillier et al. 1991 

degenerative and (Sprague- 5 dwk 

Dawley) 1x/d regenerative foci in 

tubules, accumulation of 

(GO) protein droplets in tubular 

cells) 

6 Rat 5d Endocr 50 F (decr serum thyroxine) Foster et al. 1993 

(Sprague- 1x/d 

Dawley) (GO) 

7 Rat 1 wk Hepatic 5F 16 F (incr hepatic AM-S Goldstein et at. 1978 

activity) (CD) (F) Bd Wt 159 F 0

cn 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) I 

rn 

K~Ytoa Species figure (Strain) 

8 Rat 

(Wistar) 

9 Rat 

(Wistar) 

Exposure1 

duration/ 

frequency (Specific route) System 

Id Hepatic 

(F) 

Gd 6-21, 6-1 6, Bd Wt 

6-9 or 10-1 3 

NOAEL 

(mg/k@day) 

40 F 

LOAEL 

Less serious 

(mglkglday) 

50 F (incr highly carboxylated 

porphyrins in liver) 

Serious 

(mglk Way) 

80 F (loss of body weight by 

pregnant rats) 

Reference 

Chemical Form 

Kennedy and Wigfield 1990 

Khera 1974a 

U 

A 

Lz 

B 

is

rn 

rn 

z 

10 Rat 

(SpragueDawley) 

11 Rat 

(Wistar) 

12 Rat 

(SpragueDawle y) 

Hepatic 

Hepatic 

Endocr 

Hepatic 

700 F 

250 F 

1400 F (incr ornithine 

decarboxylase in the 

liver) 

500 F (decr hepatic URO-D 

250 F (decr serum T4 levels) 

activity) 

25 F (incr urinary and hepatic 

porphyrins) 

Kitchin and Brown 

1989 

Kleiman de Pisarev et 

al. 1990 

Krishnan et al. 1991 

w 

5 

I 

rn 

T 

n 

m 

0

2 

5 

13 Rat 

(SpragueDawley) 

Hepatic 

Bd Wt 25 M 

10M (incr liver wt) Mehendale et al. 1975 

14 Rat 

(BrownNorway) 

Hepatic 45 F (increased liver weight) Michielsen et al. 2001 

15 Rat 

(W 

Bd Wt 221 M Simon et al. 1979 

P 

0 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) P 

0

I Exposure1 LOAEL 

duration/ 57J

K~Ytoa Species frequency NOAEL Less serious Serious Reference 

figure (Strain) (Specific route) System (mglkglday) Chemical Form rn (mglkglday) (mgntglday) z 

N van Raaij et al. 1993a rn 16 Rat 2wk Endocr 484 M 740 M (decr serum total and z 

rn (Wistar) 3 d/wk free T4) 

1 x/d Bd Wt 997 M 

(GW) 

Courtney et al. 1976 17 Mouse Gd7-16 Hepatic 100 F (incr liver wt in pregnant 

dams) (CD-1) (GO) 

Bd Wt 100 F 

De Matteis et al. 1961 18 GnPig 10d Hepatic 200 F (fatty change) 

(NS) (F) 

0 

200 F (incr ammonia in blood) I Renal rn

D

Neurological 5

I 

80 F (hyperesthesia, tremors, and Khera 1974a rn

19 Rat Gd 6-21, 6-1 6, 40 F n 

n 

(Wistar) 6-9 or 10-1 3 convulsions in pregnant rats) rn 

1 x/d 2 

v, 

(GO,GW) 

De Matteis et al. 1961 200 F (marked weakness, 20 GnPig 10d hyperexcitability, tremors, 

(NS) (F) convulsions) 

Reproductive 

Foster et al. 1993 50 F (incr serum 21 Rat 5d 

progesterone) (Sprague- 1 x/d 

Dawley) (GO) 

Simon et al. 1979 

22 Rat 5d 70 M 221 M (decr male impregnation 

of females) (NS) - x/d 

(GO) 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene 

~ ~~ 

- Oral (continued) 

n 

Key tofigure 

Exposure1 

duration/

Species frequency 

(Strain) (Specific route) System NOAEL 

(mglkglday) 

Less serious 

(mglkglday) 

LOAEL 

Serious 

(mglkglday) 

Reference 

Chemical Form 

Developmental 

23 Rat 

(SpragueDawley) 

4d 

Ix/d 

(GO) 

2.!jb (hyperactivity in young 

PUPS) 

3 Goldey and Taylor 

1992 

24 Rat 

(wistar) 

Gd 6-21, 6-1 6, 

6-9 or 10-13 

1x/d 

40 (incr skeletal variations) Khera 1974a 

(GO,GW) 

25 Mouse 

(CD-1) 

Gd 7-16 

(GO) 

100 (incr abnormal fetuses per 

litter) 

Courtney et al. 1976 w 

I

6

F

-I

I 

P 

N 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) I 

Exposure1 

duration/ 

INTERMEDIATE EXPOSURE 

K~Ytoa Species frequency figure (Strain) (Specific route) System 

Death 

NOAEL 

(mglkglday) 

Less serious 

(mg/kdday) 

LOAEL 

Serious 

(WJWdaY) 

Reference 

Chemical Form om 

Z 

rn 

rn 

rn 

z 

N 

26 Rat 

(Wistar) 

80d 

(F) 

100 F (60/90 died) Cuomo et al. 1991 

27 Rat 

(Wistar) 

13 wk 

(F) 

19 F (4/9 died) Den Besten et al. 1993 

28 

29 

30 

Rat 

Rat 

(SpragueDawley) 

(Wistar) 

Rat 

(Sherman) 

4 gen 

(F) 

56 d 

(F) 

4 mo 

(F) 

8 F (V20 dams died) 

50 F (2/29died) 

25 (a10 males and 14/20 

females died) 

Grant et at. 1977 

Kennedy and Wigfield 

1990 

Kimbrough and Linder 

1974 

w 

I 

rn5 

I 

rn 

n 

n 

rn2 cn 

31 Rat 15 wk 50 F (5/19 died) Koss et at. 1978 

(GO) 

32 Rat 

(coBs) 

15 wk 

(F) 

32 F (incr mortality in females) Kuiper-Goodman et al. 

1977 

33 Rat 

(SpragueDawley) 

56 d 

(F) 

100 M (13/33 died within 1 month) Ockner and Schmid 

1961 

34 Rabbit 12 wk 245 F (4/4 died) De Matteis et at. 1961 

(NS) (F) P 

W 
K~Ytoa Species figure (Strain) 

35 Pig 

(SPF) 

Exposure/

duration/ 

frequency (Specific route) 

90 d 

(F) 

Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

LOAEL 

NOAEL Less serious Serious 

System (mglkglday) (mglkglday) (mglkglday) 

50 M (5/5 died) j 

Reference 

ChemicalForm 

DenTonkelaar et al. 

1978 

I

B 

E

0 

;o 

rn 

rn 

z 

N 

z 

rn 

Systemic 

36 Monkey 90d 

(Cynomo1gus)l dd 

Hemato 10 F Foster et al. 1995a 

(C) Bd Wt 10 F 

37 Monkey 

(Rhesus) 

60d 

(GW) 

1x/d 

Resp 

Cardio 

Gastro 

128 F 

128 F 

128 F 

latropouloset al. 1976 

w 

-.? 

Musc/skel 

Hepatic 

Renal 

128 F 

8 F (hepatocellular 

hypertrophy,cloudy 

swelling) 

8 F (vacuolizationof 

proximal renal tubules) 

+i 

rn 

0

2 

Endocr 128 F 

Ocular 128 F 

38 Monkey 90d 

.. 

Hepatic 

Endocr 

Bd Wt 

0.1 F 

1F 

10 F 

1 F (hepatocellular 

vacuolation, intrahepatic 

cholestasis) 

10 F (incr adrenal weight) 

Jarrell et al. 1993 

39 Monkey 

(Rhesus) 

60d 

1 x/d 

Hemato 128 F Knauf and Hobson 

1979 

(GW) Hepatic 

Renal 

Bd Wt 

64 F 

64 F 

128 F (incr serum AST) 

128 F (incr blood urea nitrogen) 

8 F (unspecifiedweight loss) PP 
I Table 3-2. Levels of Significant Exposureto Hexachlorobenzene - Oral (continued) F 

Exposure/

duration/ 

K~Yto" Species frequency 

figure (Strain) (Specific route) 

40 Rat 5 wk 

(Fischer 344) 5 d/wk

1 dday 

System 

Hepatic 

Renal 

NOAEL 

(mglkglday) 

0.1 

25 M 

Less serious 

(mglkglday) 

1M (incr liver wt) 

LOAEL 

Serious 

(mglkglday) 

Reference 

Chemical Form 

Andrews et al. 1988 

0

I 

;o 

5 

rn 

rn 

z 

N 

z 

rn 

10M (incr serum 

1,25-dihydroxy- vitamin 

D3 and parathyroid 

hormone) 

41 Rat 15 wk 

(Fischer344) 5 d/wk 

1 x/d 

(GO) 

Musc/skel 

Hepatic 

Renal 

Endocr 

0.1 M 

1M 

1M 

0.1 M 

1M (incr femur density) 

10M (incr liver wt) 

10M (incr kidney wt, incr 

urinary LDH and alkaline 

phosphatase) 

1,25-dihydroxy- vitamin 

1 M (incrserum 

D3) 

Andrews et al. 1989 

w 

I 

rn

D

5

i 

rn 

-n 

n 

rn 

0

I 

(n 

Bd Wt 25 M 

42 Rat 15 wk 

(Fischer344) 5d/wk 

1 x/d 

(GO) 

Musc/skel 

Renal 

Endocr 

0.1 M 

0.1 M 

1M 

1M (incr femur density and 

cortical area) 

1M (incr urinary LDH) 

10M (incr serum parathyroid 

hormone) 

Andrews et al. 1990 

43 Rat 

(Wistar) 

1, 2, 3,or 4 

5 d/wk 

wks 

Hepatic 1000 F (increased porphyria, 

altered hepatic enzyme 

levels) 

Billie de Catabbi et al. 

2000a 

(GW) 

44 Rat 

(wistar) 

1 or 7 wks 

# of doses per 

wk unknown 

Hepatic 1000 F (increased urinary 

porphyrin excretion) 

Billie de Catabbi et al. 

2000a 

(GW P 

Ln 
K~Y toa Species figure (Strain) 

45 Rat 

(Wistar) 

46 Rat 

(Chbb 

THOM) 

47 Rat 

(SpragueDawley) 

48 Rat 

(Wistar) 

49 Rat 

Chbb 

THOM) 

Exposure/ 

duration/ 

frequency (Specific route) 

1x/d

5 dwk 

7, 14.21, or 28 

d 

(GW) 

1x/d 

5 dlwk 

7, 14,21, 01 

28 d 

(GW) 

7wk 

5 dlwk 

1 x/d 

(GO) 

80 d 

(F) 

1x/d 

5 dlwk 

1,2,3, 4, Or 8 

wks 

(GW) 

Table 3-2. Levels of Significant Exposure to Hexachlorobenzene 

LOAEL 

NOAEL Less serious 

System (mg/kg) ~~g/kd 

Hepatic 50 F (hepatic porphyria changes in hepatic 

sphingolipid levels) 

Hepatic 50 F (hepatic porphyria 

changes in hepatic 

sphingolipid levels) 

Renal 50 M (incr kidney wt, 

glucosuria, proteinuria, 

degenerative and 

regenerative foci in 

tubules, accumulation of 

protein droplets in tubular 

cells) 

Hepatic 100 F (incr liver wt, decr 

hepatic URO-D, incr 

hepatic porphyrins, incr 

lipid peroxidation) 

Hepatic 1000 F (Hepatic porphyria) 

- Oral (continued) 

Serious 

(mg/kg) 

Reference 

Chemical Form 

Billie de Catabbi et al. 

2000b 

Billie de Catabbi et al. 

2000b 

Bouthillier et at. 1991 

Cantoni et al. 1990 

Cochon et al. 2001 

w 

-

&

5

I 

‘ I

F

0

F 

% 

$ 

rn 

2 

N 

m 

z 

rn 

P 

0) 
K~Ytoa Species figure (Strain) 

50 Rat 

Exposure/

duration/ 

frequency (Specific route) 

1x/d 

Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

LOAEL 

(mg/kg) (mdkg) (mglkg) 

NOAEL Less serious Serious 

System 

Hepatic 1000 F (Hepatic porphyria) 

Reference 

Chemical Form 

Cochon et al. 2001 

I

F

ii 

;D 

E 

8 

L?

L 

N 

z

rn 

(wistar) 5 dlwk 

1, 2, 3,4, or 8 

rn 

wks 

(GW) 

51 Rat 

(Wistar) 

80 d 

(F) 

Hepatic 

Dermal 

100 F (porphyria, incr liver wt, 

fatty degeneration) 

100 F (photosensitive skin 

Cuomo et al. 1991 

lesions) 

Bd Wt 100 F w 

52 Rat 

(Wistar) 

13 wk 

(F) 

Hepatic 9.5 F (incr liver wt, incr urinary 

and liver porphyrins) 

Den Besten et al. 1993 $ 

I

5 

Renal 9.5 F 19 F (incr kidney wt, 

basophilic renal tubules, 

protein casts) 

rn 

n 

n 

rn 

0 

-1 

Endocr 9.5 F 19 F (adrenal: incr wt, cortical 0, 

hypertrophy and 

hyperplasia 

thyroid: decr serum T4 

and T3) 

Dermal 9.5 F (skin lesions) 

Bd Wt 9.5 F 19 F (severe wt loss in 419) 

53 Rat 

(Wistar) 

107 d 

(F) 

Hepatic 150 F (marked decr in hepatic 

URO-Dactivity, massive 

incr in hepatic porphyrin 

Elder and Urquhart 

1986 

content) 

54 Rat 

(wistar) 

4wk 

1 x/d 

5 dlwk 

Renal 1000 F (porphyria and lipid 

peroxidation in cortex) 

Fernandez-Tome et al. 

2000 

(GW) 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

Exposure/

duration/ 

K~Ytoa Species frequency 

figure (Strain) (Specific route) System 

NOAEL 

(mg/kg/day) 

Less serious 

(mglkglday) 

LOAEL 

Serious 

(mglkglday) 

Reference 

Chemical Form 

55 Rat 

(Sprague30 d 1 Ad 

Endocr 1 F (decrserum 

corticosterone) 

Foster et al. 1995b 

Dawley) (GO) Bd Wt 100 F 

56 Rat 

(CD) 

4 mo 

(0 

Resp 4F 12 F (hypertrophy and 

proliferationof 

macrophages) 

endothelial cells, incr 

Goldstein et al. 1978 

Cardio 36 F 

Hepatic 

Endocr 

4F 

36 F 

12 F (incr urinary and hepatic 

porphyrins, enlarged 

hepatocytes) 

w 

I 

rn

5

I 

57 Rat daily 

(GO) 

(Fischer- 344)s wks 

Dermal 

Bd Wt 

Hepatic 

12 F 

36 F 

36 F (small sores) 

28 M (liver enlargement, 

porphyria, increased 

enzyme expression, 

increased protooncogene 

expression) 

Gustafson et al. 2000 

rn 

-n 

n 

rn 

v) 

2 

58 Rat 

(Wistar) 

56 d 

(F) 

Hepatic 

Renal 

5 F (incr highly carboxylated 

porphyrins in liver) 

5 F (incr highly carboxylated 

porphyrins in kidney) 

Kennedy and Wigfield 

1990 

Bd Wt 5F 50 F (large decr body wt gain) 

P 

m 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

Exposure1

duration/ 

LOAEL 

K~Ytoa Species frequency NOAEL Less serious Serious Reference 

figure (Strain) (Specific route) System (mdkglday) (mglkglday) (mglkglday) Chemical Form 

59 Rat 

(Sherman) 

4 mo 

(F) 

Resp 5 (incr macrophages, focal 

areas of fibrosis) 

25 F (extensive intra-alveolar 

hemorrhage, edema in 

females) 

Kimbrough and Linder 

1974 

Cardio 5 25 (fibrosis, degeneration) 

Hemato 5 (decr hemoglobin and 

hematocrit) 

Hepatic 5 (incr liver wt, enlarged 25 (necrosis, fibrosis) 

hepatocytes) 

Renal 5 25 (incr kidney wt) 

60' Rat 

(SpragueDawley) 

1 gen 

(F) 

Endocr 

Dermal 

Bd Wt 

Resp 

5 

5 

5 (hyperplasia of the 

adrenal cortex) 

25 (skin eruptions) 

25 (decr body wt gain in 

males) 

3 F (intraalveolar foamy 

histiocytes, hypertrophy 

and proliferation of 

endothelial cells of 

Kitchinet at. 1982 

w 

I

W 

-I

I 

F 

rn 

n 

n

W 

pulmonary venules in 

dams) 

Bd Wt 7F 

61 Rat 

(Wistar) 

4 wk 

(GW) 

1 x/d 

Endocr 1000 F (decr serum T4, incr T4 

metabolism, incr serum 

TSH, incr thyroid iodine 

uptake) 

Kleiman de Pisarev et 

al. 1989 

Bd Wt 1000 F 

62 Rat 

(Wistar) 

30 d 

1 x/d 

Endocr 125 F 500 F (decr serum T4 levels) Kleiman de Pisarev et 

at. 1989 

(GW) 

P 

W 
Exposure1 

P duration1 

Key to" Species frequency figure (Strain) (Specific route) 

63 

64 

65 

66 

67 

Rat 

(Wistar) 

Rat 

(Wistar) 

Rat 

(Wistar) 

Rat 

(SpragueDawley) 

Rat 

(SpragueDawley) 

8wk 

7 dlwk 

1 xld 

(GW) 

4 wk 

7 dwk 

1xld 

(GW) 

15 wk 

3-4 dlwk 

1 xld 

(GO) 

3-6 wk 

5 dlwk 

1xld 

(GO) 

6wk 

5 dlwk 

1xld 

(GO) 

Table 3-2. Levels of Significant Exposure to Hexachlorobentene - Oral (continued) 

LOAEL 

System 

NOAEL 

(mg/kg/day) 

Less serious 

(mglkglday) 

Serious 

(mglkglday) 

Hepatic 1000 F (incr liver wt, decr 

URO-D activity, incr 

hepatic and urinary 

porphyrins, incr 

dehalogenation of T4) 

Endocr 1000 F (decr serum T4, incr 

serum TSH) 

Hepatic 1000 F (incr liver wt) 

Endocr 1000 F (decr serum T4, incr 

serum TSH) 

Bd Wt 1000 F 

Hepatic 50 F (incr liver wt, massive 

incr liver and urine 

porphyrins) 

Renal 50 F (incr kidney wt) 

Endocr 50 F (incr adrenal wt) 

Dermal 50 F (rough appearance of 

fur, hair loss, skin 

lesions) 

Bd Wt 50 F 

Hepatic 50 (incr urinary and hepatic 

porphyrins) 

Hepatic 50 F (incr hepatic and urinary 

porphyrins) 

Reference 

Chemical Form 

Kleimande Pisarev et 

al. 1990 

Kleirnan de Pisarev et 

al. 1995 

Koss et al. 1978 

Krishnan et al. 1991 

Krishnan et al. 1992 

-I

I 

rn 

I

?

i-,

I 

i;

% 

rn 

z 

N 

rn 

z 

rn 

ul 

0 
Exposure1

duration/ 

K~Ytoa Species frequency figure (Strain) (Specific route) 

68 Rat 15 wk 

(COBS) (F) 

Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

LOAEL 

Less serious Serious 

(mglkglday) 

NOAEL 

System (mwdday) 

Hemato 8 32 F (decr red blood cell count, 

(mglkglday) 

hernatocrit, and 

hemoglobin; leukocytosis 

in females) 

Reference 

Kuiper-Goodman et al. 

Chemical Form 

1977 

I

F 

E

0 

A 

rn 

rn 

rn 

z 

N 

z 

Musc/skel 

Hepatic 

32 

0.5 2 (basophilic clumping in 

hepatocytes) 

Renal 8 32 (incr kidney wt) 

69 Rat 

(Wistar) 

30 d 

(F) 

Dermal 

Bd Wt 

Resp 

8 

8 

32 (alopecia) 

32 M (1 3% decr in terminal 

body wt in males) 

22 F (cuboidal venule 

endothelial cells, 

perivascular infiltrate, 

accumulationof alveolar 

macrophages) 

Michielsen et al. 1997 

w 

I 

F 

5

i 

rn

9 

rn 

7 

rn 

Hepatic 

hypertrophy with 

cytoplasmic inclusions) 

22 F (incr liver wt, hepatocyte 

Renal 45 F 

Dermal 22 F (gross skin lesions, 

epidermal hyperplasia, 

inflammatory infiltrate in 

the dermis) 

Bd Wt 45 F 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

K~Ytoa Species 

Exposure/

duration/ 

frequency NOAEL Less serious 

LOAEL 

Serious Reference 

figure (Strain) (Specific route) System (mglkglday) (mglkglday) (mglkglday) Chemical Form 

70 Rat 

(Lewis) 

29 d 

(F) 

Resp 8 F (cuboidal venule 

endothelial cells, 

accumulation of alveolar 

macrophages) 

perivascular infiltrate, 

Michielsen et al. 1997 

Hepatic 8 F (incr liver wt, hepatocyte 

hypertrophy with 

cytoplasmic inclusions) 

Renal 22 F 

71 Rat 

(Brown 

Norway) 

28 d 

(F) 

Dermal 

Bd Wt 

Resp 

22 F 

8 F (gross skin lesions, 

epidermal hyperplasia, 

inflammatory infiltrate in 

the dermis) 

8 F (cuboidal venule 

endothelial cells, 

perivascular infiltrate, 

accumulation of alveolar 

macrophages) 

Michielsen et al. 1997 

w 

I 

rn

5

I 

rn

9 

rn

3 

Hepatic 

hypertrophy with 

cytoplasmic inclusions) 

8 F (incr liver wt, hepatocyte 

Renal 22 F 

Dermal 8 F (gross skin lesions, 

epidermal hyperplasia, 

inflammatory infiltrate in 

the dermis) 

Bd Wt 22 F 45 F (sudden weight loss) 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) I 

s 

Exposure1 LOAEL 9 

duration/ 5 K~Ytoa Species frequency NOAEL Less serious Serious Reference z1 

0 figure (Strain) (Specific route) System (,"glkglday) (mglkglday) (mglkdday) Chemical Form m

!!? 

Michielsen et al. 2000 is 72 Rat daily Hepatic 67 F (increased liver weight) rn 

(Brown- 4 wks rn 

Z 

Norway) (F) Resp 67 F 

Dermal 67 F (dermal lesionshyperplasia, deep 

venules with activated 

endothelium, 

inflammation) 

Michielsen et al. 2001 73 Rat 7 or 21 d Hepatic 45 F (increased liver weight) w 

(Brown- daily 

Norway) (GW) Dermal 45 F (lesions with inflammatory 

mg/kg response) rn 

n Resp 45 F (airway 

2 

n 

rn mg/kg hyperresponsiveness, 

granuloma (non-neoplastic cn 

lesions) formation) 

74 Rat 7or21 d Hemato 45 F (increased total IgE) Michielsen et al. 2001 

(Brown- daily 

Norway) (GW) Dermal 45 F (lesions, inflammatory 

mg/kg cellular imune response) 

Resp 45 F (airway 

mg/kg hyperresponsiveness, 

granuloma (non-neoplastic 

lesions) formation ) 

75 Rat 60 d Hepatic 430 M (incr serum ALT, Nikolaev et al. 1986 

bilirubin, and glutamate (Wistar) (F) dehydrogenase) 

(n

W 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

I 

76 Rat 

(SpragueDawley) 

Exposure1

duration/ 

K~Ytoa Species frequency 

figure (Strain) (Specific route) 

56 d 

Musclskel 

Hemato 

System 

NOAEL 

100 M 

(mdkdday) 

LOAEL 

Less serious 

(mglkglday) 

100M (porphyrin accumulation 

in cortex of long bones) 

Serious 

(mglkdday) 

Reference 

Chemical Form 

Ockner and Schrnid 

1961 

I

i 

m 

m 

rn 

z 

N 

z 

Hepatic 100M (porphyria, 

hepatomegaly, 

hepatocellular 

degeneration) 

77 Rat 

(Wistar) 

Hepatic 

Dermal 

Bd Wt 50 M 

25 M (incr liver wt) 

25 M (transient lesions on 

face and ears) 

Schielen et al. 1993 I*' 

I 

rn

5

I 

m 

-n 

n 

78 Rat Dermal 50 (skin lesions) Schielen et al. 1995b rno 

(Wistar) 

Bd Wt 50 

79 Rat 

(Agus) 

Hepatic 5 F (incr hepatic porphyrin 

content, decr hepatic 

URO-D activity, incr 

hepatic AM-S activity) 

Smith et al. 1979 

80 Rat Hepatic 5F Smith et al. 1979 

(PortonWistar) 

vl

P 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) I 

~~ 

Exposure1

duration/ 

K~Ytoa Species frequency 

figure (Strain) (Specific route) 

81 Rat 15 wk 

(Fischer- 344) (F) 

System 

Hepatic 

NOAEL 

(mS/k@day) 

LOAEL 

Less serious 

10 (incr liver wt, 

(mglkglday) 

hepatocellular 

hypertrophy, incr hepatic 

porphyrin levels in 

females) 

Serious 

(mS/kCWy) 

Reference 

Chemical Form 

Smith et al. 1985 

0

I 

73 

rn

z 

N 

Z 

rn 

6 

rn 

Renal 10 F (incr renal porphyrin 

levels in females) 

82 

83 

Rat 

(Wistar) 

Rat 

(Wistar) 

Hepatic 

Endocr 

Hepatic 8M 

1000 F (decr URO-D activity in 

liver) 

1000 F (decr serum T4, incr 

serum TSH) 

22 M (incr liver wt) 

Sopena de Kracoffet 

al. 1994 

Van Loveren et at. 

1990 

F 

84 Rat 

(Wistar) 

Hepatic 

Endocr 

1000M (incr liver wt) 

1000M (decr serum T4 levels) 

van Raaij et al. 1993b rn

2 

cn 

85 Rat 

(Wistar) 

Hemato 25 M (incr extramedullary 

hematopoiesis in spleen; 

neutrophilia) 

Vos et al. 1979b 

Hepatic 25 M (incr liver wt, liver Cell 

hypertrophy and 

cytoplasmic 

hyalinization) 

Renal 100 M 

Endocr 100 M 

Bd Wt 100 M 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) I 

Exposure1 

duration/ 

Key toa Species frequency

figure (Strain) (Specific route) 

86 Rat 

(BD VI) 

System 

Hepatic 

NOAEL 

(mg/kg/day) 

LOAEL 

Less serious 

(mglkglday) 

1000 (decr URO-D activity in 

males and females; incr 

hepatic and urinary 

porphyrins in females) 

Serious 

(mglkglday) 

Reference 

Chemical Form 

Wainstok de 

Calmanovici et al. 

1991 

;o 

mis

rn 

rn 

Z 

87 Mouse 

(Balb/c) 

Resp 22 M Loose et at. 1977 

88 

89 

Mouse 

(C57BU 

1OScSn) 

Dog 

(Beagle) 

Hepatic 

Hepatic 

Resp 

Cardio 

Hemato 

100 F 

100 F 

100 F 

22 M (incr liver wt, 

hepatocellular 

hypertrophy) 

26 M (incr hepatic porphyrin 

levels) 

Vincent et at. 1989 

Sundlof et al. 1981 

wL 

m

F

-I

I 

rn 

7 

m 

v) 

2 

Hepatic 50 F (fatty changes in liver, 

swollen hepatocytes and 

hepatomegaly) 

Renal 100 F 

Endocr 100 F 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) I

P

0 Exposure1 LOAEL 

duration/ z 

K~Y frequency Reference toa Species ;a NOAEL Less serious Serious 

figure (Strain) (Specific route) System (mgkgday) (mglkglday) (mglkglday) Chemical Form rn 

Z 

N De Matteis et al. 1961 90 Rabbit 12 wk Hemato 245 F rn

z 

rn

(NS) (F) Musc/skel 245 F (necrosis, degeneration, and 

focal calcification in muscle, 

incr porphyrins in bone) 

Hepatic 245 F (incr urinary and hepatic 

porphyrins, fatty change, 

necrosis) 

Bd Wt 245 F (weight loss) u 

91 Pig 90 d Hemato 50 M Den Tonkelaar et al. I 

rn 1978 

(SW (F) 5

Hepatic 0.05 M 0.5 M (hepatocellular I

rn hypertrophy) n

n 

2

rn Renal 0.5 M 5 M (incr kidney wt) 50 M (degeneration of proximal 

tubules and loop of Henle) (I) 

Endocr 0.5 M 5 fvl (incr thyroid wt) 

Bd Wt 5M 50 M (depressed growth) 

Imrnunological/Lymphoreticular 

92 Monkey 60d 8 F (thymic cortical atrophy) latropoulos et al. 1976 

(Rhesus) (GW) 

93 Rat 56 d 5F 50 F (incr spleen weight, incr Kennedy and Wigfield 

1990 (Wistar) highly carboxylated (F) porphyrins in spleen) 

50 F (incr spleen wt) Koss et al. 1978 
Exposure/

duration/ 

K~Ytoa Species frequency 

figure (Strain) (Specific route) 

95 Rat 15 wk 

(COBS) (F) 

Table 3-2.Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

LOAEL 

NOAEL Less serious Serious 

System (mglkglday) (mglkglday) (mglkglday) 

8 32 (incr lymphocyte count, 

incr spleen wt, 

congestive splenomegaly 

in females) 

Reference 

Chemical Form 

Kuiper-Goodman et al. 

1977 

I

F 

z 

0 

;o 

rn 

rn 

m 

8 

z 

N 

Z 

96 Rat 

(Wistar) 

30 d 

(F) 

22 F (incr spleen wt, “high 

endothelial” venules in 

popliteal lymph node, 

incr serum IgM) 

Michielsen et al. 1997 

97 

98 

Rat 

(Lewis) 

Rat 

(Brown 

Norway) 

29 d 

(F) 

28 d 

(F) 

8 F (“high endothelial” 

venules in popliteal 

lymph node) 

8 F (incr spleen wt, “high 

endothelial” venules in 

popliteal lymph node; 

incr serum IgM 

autoantibodies) 

Michielsen et al. 1997 

Michielsen et al. 1997 

I 

m 

5

F 

I 

m

z 

rn

3 

vj 

99 Rat 

(BrownNorway) 

daily 

(F) 

4 wks 

67 F (increased spleen weight) Michielsenet al. 2000 

100 Rat 

(BrownNoway) 

7or21 d 

daily 

(GW) 

45 F (Increased spleen 

weight) 

Michielsen et al. 2001 

101 Rat 3wk 25 M (incr IgM autoantibodies) Schielen et al. 1993 

(Wistar) (F) 

102 Rat 

(Wistar) 

13 wk 

(F) 

8 F (incr spleen wt, incr total 

serum IgM, incr serum 

IgM autoantibodies) 

Schielen et al. 1995a 
Exposure1

duration/ 

K~Ytoa Species figure (Strain) frequency (Specific route) 

103 Rat 3wk 

(Wistar) (F) 

104 Rat 6 wk 

(Wistar) (F) 

105 Rat 3wk 

(Wistar) (F) 

106Mouse 6wk 

(F) 

107Mouse 6wk 

(Balb/c) (F) 

108Mouse 6 wk 

(Balb/c) (F) 

109 Mouse 18 wk 

(Balb/c) (F) 

Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

LOAEL 

NOAEL Less serious Serious 

System (mg/kg/day) (mglkglday) (mglkglday) 

50 (incr spleen and lymph 

node wt, altered size 

distribution of 

splenocytes, selective 

activation of splenic B-1 

cells) 

8 M (decr NK cell activity in 

lung) 

25 M (incr wt and proliferation 

of high-endothelial 

venules in lymph nodes, 

enlarged white pulp 

[marginal zones and 

follicles] in spleen, incr 

neutrophil count) 

22 (incr susceptibility to 

hepatitis infection) 

22 M (immunosuppression, 

decr antibody 

production) 

22 M (incr susceptibility to 

bacterial endotoxin and 

protozoan infection) 

22 M 

Reference 

Chemical Form 

Schielen et al. 1995b 

Van Loveren et al. 

1990 

Vos et al. 1979b w 

I 

rn

5

I 

rn 

n 

n 

rn 

0 

-I 

v) 

Carthew et al. 1990 

Loose et al. 1977 

Loose et al. 1978 

Loose et al. 1981 

cn 

W 
I Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

K~Ytoa Species figure (Strain) 

110Mouse 

Exposure/

duration/ 

frequency (Specific route) 

18wk 

(F) 

System 

NOAEL 

(mglkglday) 

tumor challenge in vivo, 

decr killing of tumor cells 

in vitro) 

LOAEL 

Less serious 

(mglkglday) 

0.6 M (incr susceptibility to 

Serious 

(mflglday) 

Reference 

Chemical Form 

Loose et al. 1981 

h 

A 

rn 

rn 

rn 

E 

z 

N 

z 

22 M (decr graft-host activity) Silkworth and Loose 

1981 

5M 50 M (atrophy of lymph 

nodes) 

Den Tonkelaar et al. 

1978 

Neurological 

1 13 Monkey 

(Rhesus) 

114 Rat 

(Wistar) 

60 d 

1 x/d 

(GW) 

1x/d 

5 dlwk 

1,2, 3, 4, or 8 

wks 

32 F 

1000 F (altered phospholipid 

levels in brain) 

64 F (severe tremors, muscular 

weakness) 

Knauf and Hobson 

1979 

Cochon et al. 2001 

I*' 

I 

m

5

I 

rn 

-n 

n 

rn 

20 

(GW) 

115 Rat 

(Chbb 

THOM) 

1 x/d 

5 dlwk 

1, 2, 3,4, or 8 

wks 

1000 F (altered phospholipid 

levels in brain) 

Cochon et al. 2001 

(GW) 

4 mo 12 F 36 F (excessive irritability) Goldstein et al. 1978 

(F) 

117Rat 

(SpragueDawley) 

4 gen 

(F) 

4 8 F (convulsions in dams) Grant et al. 1977 
Exposure1

duration/ 

K~Ytoa Species frequency 

figure (Strain) (Specific route) 

Table 3-2.Levels of Significant Exposure to Hexachlorobentene - Oral (continued) 

LOAEL 

Serious 

System NOAEL Less serious 

(mglkglday) (mglkglday) (mglkglday) 

Reference 

Chemical Form 

I 

!2

0

I

6 

;o 

rn 

118Rat 

(Wistar) 

56 d 

(F) 

5F 50 F (lethargy, tremor, 

convulsions) 

Kennedy and Wigfield 

1990 

Z 

N

rn 

z 

rn 

119Rat 

(Sherman) 

4 mo 

(F) 

5 25 (tremor, hyperexcitability) Kimbrough and Linder 

1974 

120 Rat 15 wk 50 F (muscle fasciculations, Koss et al. 1978 

(Wistar) 3-4 d/wk

1 x/d 

tremors) 

(GO) w 

121 Rat 15wk 

(F) 

8 32 (tremors, ataxia, hind limb 

paralysis) Kuiper-Goodman et al. 

1977 

I 

rn5

I 

122 Rat 

(Wistar) 

60 d 

(F) 

hyperexcitability) 

430 M (clonic convulsions, tremors, Nikolaev et at. 1986 rn 

qrn

7 

v) 

123 Rat 

(SpragueDawley) 

56 d 

(F) 

100 M (ataxia, tremor, paralysis) Ockner and Schmid 

1961 

124 Rat 20 wk 40 M (decr nerve conduction Sufit et at. 1986 

(SpragueDawley) (F) velocity) 

125 Dog 

(Beagle) 

21 d 

lor 2 x/d 

50 F (dysrhythmic electroencephalogram) 

Sundlof et al. 1981 

(C) 

126 Rabbit 12 wk 245 F (tremor, paresis) De Matteis et al. 1961 

(NS) (F) 

127Pig 90 d 5M 50 M (tremors, unsteady gait) Den Tonkelaar et al. 

1978 

(SPF) (F) 
Key tofigure 

Exposure1 

duration/ 

Species frequency

(Strain) (Specific route) 

Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

LOAEL 

NOAEL Less serious Serious

System (mglkglday) (mglkglday) (mglkglday) 

Reference 

Chemical Form 

I 

!! 

0

I 

c

0

2J

8 

rn 

Reproductive 

128 Monkey 90 d 

(Cynornolgus)1x/d 

(C) 

0.1 F (cellular degeneration of 

ovarian surface 

epithelium) 

129 Monkey 90 d 

(Cynorno1gus)l x/d 

(C) 

0.01 F (mitochondria1 

degeneration in 

developing ovarian 

follicles) 

Bourque et al. 1995 

130 Monkey 13 wk 

(C) 

131 Monkey 90 d 

(cynorno1gus)l x/d 

(C) 

0.1 F 

1F 

1 F (decrserum 

progesterone levels 

during the luteal phase 

of menstruation) 

10 F (incr length of menstrual 

cycle, decr ovulatory 

levels of estradiol) 

Fosteret al. 1992a 

Foster et al. 1995a 

-F

I 

rn

F

4

I 

rn 

rn 

0 

-I 

0) 

132 Monkey 

(Rhesus) 

60 d 

(GW) 

8 F (degeneration of the 

germinal epithelium, 

reduced number of 

primary follicles, and 

multiple follicular cysts in 

the ovaries) 

latropoulos et al. 1976 

133 Monkey 90 d 

(Cynornolgus)1x/d 

0.1 F (degenerative lesions in 

oocytes) 

Jarrell et al. 1993 

(C) 

134 Monkey 

(Rhesus) 

24 d 

1 x/d 

4 F (blocked ovulation) Muller et al. 1978 

(GW) 

m 

h, 
K~Ytoa 

figure Species(Strain) 

Exposure1 

duration1 

frequency (Specific route) 

Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

LOAEL 

NOAEL Less serious Serious 

(mglkg/day) (mdkg/day) (mglkglday) System 

~ ~ 

0.1 F (altered morphology of 

ovary surface 

epithelium cells) epithelium cells, denuding of 

1 F (necrosis of ovary surface 

ovary) 

Reference 

Chemical Form 

Sims et at. 1991 

>

8 

8

5 

rn

2 

rn 

rn 

N 

Z 

1000 F (incr estrus duration, 

altered hormone levels, 

reduced ovulation, 

degenerative ovarian 

lesions) 

Alvarez et al. 2000 

137 Rat 

(SpragueDawley) 

138 Rat 

139 Rat 

(SpragueDawley) 

(SpragueDawley) 

21 d 

1 x/d 

(GO) 

4 gen 

(F) 

daily 

28d 

pre-mating- weaning 

Bd Wt 

8 

0.008 F 

1 F (incr serum progesterone 

levels) 

16 (decr fertility and litter size) 

Foster et al. 1992b 

Grant et al. 1977 

Nakashirna et al. 1997 

w 

I 

rn

5

I 

rn 

z 

rn 

0 

-1 

v, 

(GO) 

140 Pig 

(SW 

90 d 

(F) 

5M 50 M (retarded development 

of the testes) 

Den Tonkelaar et al. 

1978 

Developmental 

141 Monkey 

(Rhesus) 

22-60 d 

1 x/d 

(GW) 

64 (2/3 infants died with 

lethargy, ataxia, and 

listlessness) 

latropouloset at. 1978 
Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) I 

K~Ytoa Species figure (Strain) 

142 Rat 

Exposure1

duration/ 

frequency (Specific route) 

4 aen -

LOAEL 

NOAEL Less serious Serious 

System (m@k&fay) (m@!$day) (mglkglday) 

2 4 (decr pup weight gain) 8 (decr pup viability) 

Reference 

Chemical Form 

Grant et al. 1977 

C J I 

i;70 

rn 

z 

N 

(SpragueDawley) (F)

L 

rn 

143 Rat 1 gen 3 (decr pup weight gain) 4 (decreased survival) Kitchin et al. 1982 

(SpragueDawley) (F)

144 Rat 2 gen 0.6 M 1.3 M (reduced efficiency of Lilienthal et al. 1996 

(Wistar) (F) pups in operant behavior 

task) 

145 Rat 

(Wistar) 

gestation + 

lactation + 

2 wks 

post-weaning 

(F) 

2.5 (decr resistance to 

infection, incr IgG 

response to tetanus 

toxoid, and proliferation 

of high endothelial 

venules in lymph nodes 

in pups) 

Vos et al. 1979a w 

I 

rn

Fi

I 

rn 

n 

n 

rn 

146 Rat 

(wistar) 

gestation + 

lactation + 

0.2 (incr IgG and IgM 5 (high pup mortality) 

responseto tetanus 

Vos et al. 1983 

2 wks or toxoid, incr delayed-type 

7 mo hypersensitivity reaction 

post-weaning to ovalbumin, and 

(F) 

accumulation of foamy 

macrophages in the lung 

in offspring) 

cn 

P 
duration/ 

Exposure1 

K~Ytoa Species frequency 

figure (Strain) (Specific route) 

CHRONIC EXPOSURE 

Death 

Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

LOAEL 

NOAEL Less serious Serious 

System (mglkglday) (mglkglday) (mglkglday) 

Reference 

Chemical Form 

I

x 

0

F

0 

;D 

rn 

rn 

rn 

z

N 

z 

147 Mouse 120 wk 24 (decr survival) Cabral et al. 1979 

(Swiss) (F) 

148Hamster 

(Syrian) 

lifespan 

(F) 

16 (decr lifespan) Cabral et al. 1977 

149 Dog 

(Beagle) 

Systemic 

150 Rat 

(SpragueDawley) 

1 Yr 

1dd 

(C) 

2 gen 

(F) 

Hemato 

Hepatic 

2 

0.08 0.4 M (incr liver wt) 

11 F (2/6 females died) Gralla et at. 1977 

Arnold et al. 1985 

w 

I 

rn

5

I 

rn 

-n 

rn 

0 

--I 

0) 

Renal 2 

Bd Wt 2 

151 Rat daily 

(Fischer-344)s wks 

(F) 

Hepatic 6.25 6.25 (increased liver weight, 

centrilobular 

hypertrophy,foci of 

GST-P) 

Kishirnaet al. 2000 

152 Rat 

(SpragueDawley) 

1 Yr 

(F) 

Hepatic 0.05 0.25 (mitochondria1swelling 

and elongation) 

Mollenhaueret al. 

1975 

153 Rat 90 wk Hepatic 7 F (porphyria) Smith and Cabrall980 

(Agus) (F) Dermal 

Bd Wt 

7 F (alopecia) 

7 F (decr body wt gain) ln

Ln 
Exposure1

duration/ 

K~Ytoa Species frequency 

figure (Strain) (Specific route) 

154 Rat 90wk 

(Fischer- 344) (F) 

Table 3-2.Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

LOAEL 

NOAEL Serious Less serious System (,,,@&$day) (mg/kg/day) (mglkglday) 

Hepatic 10 (decr URO-D activity, 

incr hepatic porphyrin 

levels, hepatocyte 

hypertrophy, fatty 

degeneration, bile duct 

hyperplasia) 

Reference 

Chemical Form 

Smith et al. 1985 

I

P

0

F

0 

?J 

8

r? 

i? 

m 

z 

rn 

Renal 10 (incr kidney wt, incr 

renal porphyrin levels, 

nephrosis in males) 

155 Rat 65 wk 

(Fischer- 344) (F) 

Endocr 

Bd Wt 

Hepatic 

10 

10 (decr body wt) 

10 F (incr liver wt; biliary 

hyperplasia; incr liver 

porphyrins; induction of 

microsomal enzymes 

and glutathione 

S-transferase) 

Smith et at. 1993 

I*’ 

I 

rn

5

I 

m 

n 

n 

rn

2 

(I) 

Bd Wt 10 F 

156 Mouse 18mo Hepatic 13M (hepatocyte 

hypertrophy) 

Smith et at. 1989 

Bd Wt 13 M (decr body wt) 

157Hamster lifespan Bd Wt 16 M (marked decr in wt gain) Cabral et al. 1977 

(Syrian (F) golden) 
Exposurd 

duration/ 

K~Y toa Species frequency figure (Strain) (Specific route) 

158 Dog 1Yr 

(Beagle) (C) 

Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

LOAEL 

NOAEL Less serious Serious System (mglkglday) (mglkglday) (mglkglday) 

Cardio 11 110 (arteriopathy) 

Gastro 1 11 (diarrhea, necrotic and 

inflammatory lesions of the 

omentum and abdominal 

Reference 

Chemical Form 

Gralla et at. 1977a 

I

E 

0 

E

7J

8 

rn 

z 

m

rn 

z

rn

serosa) 

Hemato 1 11 (neutrophilia) 110 (anemia) 

Hepatic 

Bd Wt 

1 

1 

11 (hepatomegaly, bile duct 

hyperplasia, pericholangitis, 

periportal fibrosis, incr serum 

alkaline phosphatase and 

AST) 

11 (loss of body weight) 

0 

I 

rn 

5

I 

lmmunologicallLymphoreticular rn

n 

n 

159 Dog 1Yr 

(Beagle) (C) 

0.1 (incr severity of nodular 

hyperplasia of the gastric 

lymphoid tissue) 

Gralla et al. 1977a 2rn 

(n 

Neurological 

160Mouse 120wk 24 (tremors, convulsions) Cabral et al. 1979

(Swiss) (F) 

Reproductive 

161 Rat 2 gen 2 Arnold et al. 1985 

(Sprague- (F) Dawley) 
Exposure/

duration/ 

K~Ytoa Species frequency figure (Strain) (Specific route) 

Developmental 

162 Rat 2 gen 

(F) (SpragueDawley) 

Table 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

Less serious Serious Reference 

(mglkglday) (mglkglday) 

LOAEL 

NOAEL 

System @,glkglday) Chemical Form 

0.016d M (peribiliary lymphocytosis 2 

and fibrosis of the liver in 

F1 adult males) 

(decr neonatal viability of F1 Arnold et al. 1985 

pups; incr neoplastic liver 

nodules, parathyroid 

adenoma, and adrenal 

pheochromocytoma in F1 

adults) 

I

P

0 

0 

T 

;D

8 

rn 

z 

N 

Z 

rn 

m 

Cancer 

163 Rat 

(SpragueDawley) 

164 Rat 

(Ag4 

165 Rat 

(Wistar) 

104 wk 

(F) 

90 wk 

(F) 

75 wk 

(F) 

4 

7F 

5F 

(CEL: hepatoma, bile-duct 

adenoma, hepatocarcinoma, 

renal adenoma) 

ErtDrk et al. 1986 

Smith and Cabral 1980 (CEL: liver-cell tumors) 

(CEL: liver-cell tumors) Smith and Cabrall980 

w 

I 

rn

5

I 

$I 

n 

rn 

v, 

9 

166 Rat 90 wk 

(Fischer- 344) (F) 

10 (CEL: hepatocarcinoma) Smith et al. 1985 

167 Rat 65 wk 

(Fischer- 344) (F) 

10 F (CEL: liver tumors) Smith et al. 1993 

168Mouse 

(Swiss) 

120 wk 

(F) 

12 (CEL: liver tumors) Cabral et al. 1979 
Exposure/ 

duration/ 

Key to- Species frequency 

figure (Strain) (Specific route) 

169 Hamster lifespan 

(Syrian) (F) 

Table 3-2.Levels of Significant Exposure to Hexachlorobenzene - Oral (continued) 

LOAEL 

NOAEL Less serious Serious 

System (mglkglday) (mglkglday) (mglkglday) 

4 (CEL: hepatoma, liver 

hemangioendothelioma, 

thyroid alveolar adenonla) 

Reference 

Chemical Form 

Cabral et al. 1977 

I 

m X

R

I

T-T

V

R v

W

e! 

is 

rn 

z 

rn 

'The number corresponds to entries in Figure 3-2. 

to derive an acute oral minimal risk level (MRL) of 0.008 mg/kg/day; dose divided by uncertainty factor of 300 (10 for extrapolation from rats to humans, 10for human 

variability, and 3 for use of a minimal LOAEL). 

'Used to derive an intermediate oral MRL of 0.0001 mglkglday; dose divided by an uncertainty factor of 100 (3 for extrapolation from monkeys to humans, 10for human 

variability, and 3 for use of a minimal LOAEL). 

dUsed to derive a chronic oral MRL of 0.00005 mgkglday; dose divided by an uncertainty factor of 300 (10 for extrapolation from rats to humans, 10for human variability, and 3 

for use of a minimal LOAEL). w 

ALA-S = delta-aminolevulinic acid synthetase; AST = aspartate aminotransferase; Bd wt = body weight; (C) = capsule; Cardio = cardiovascular; CEL =cancer effect level; d = 

day(s); CNA = central nervous system; Decr = decreased; Endocr = endocrine; F = female; (F) = feed; (G) = gavage; Gastro = gastrointestinal; Gd = gestational day; gen = 

generation; (GO) = gavage in oil; (GW) = gavage in water; Hemato = hematological; Incr = increased; LD,, = lethal dose, 50% kill; LDH = lactose dehydrogenase; LOAEL = 

I

rn 

5

I 

lowest-observable-adverse effect level; M = male; mo = month(s); Muschkel = musculoskeletal; NOAEL = no-observable-adverse-effect level; Resp = respiratory; T3 = 

tniodothyronine; T4 = thyroxine; TSH = thyroid stimulating hormone; URO-D = uroporphyrinogen decarboxylase; wk = week(s); wt = weight; x = times; yr = year(s) 

rn 

n 

n

rn

2 

v, 
Figure 3-2. Levels of Significant Exposure to Hexachlorobenzene- Oral 

Acute (114 days) 

1oooc -

1OO( 

10' 

1 

-

-

Ollr @lag 

017rn 

@14r 

Q12r 

Q7r 

a1 3r 

O7r 

a189 

Q5r 

0209 

019r 

019r 

025m 

Q24r 

Q23r 

rn 

rn 

(I) 

n 

n 

2 

u 

I 

rn

5

I 

0. 

0.0 

0.00I 

-
Figure 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (Continued) 

Intermediate (15-364 days) 

Systemic 

mglkglday e@ 

10000 

1000 

034h 090h 

100 026r 033r 037k O89d 

3%: 

037k 089d 039k 076r

8;;: 09lp Q85r w 

10 028r 

a56r 

@59r 

@70r @71r 

059r 

036k 068r 

@59r 

056r @60r 

1 

0.1 

0.01 

0.001 

0.0001 

c-Cat -Humans f-Ferret n-Mink Cancer Effect Level-Animals V Cancer Effect Level-Humans LD50/LC50

d-Dog k-Monkey j-Pigeon o-Other 0LOAEL, More Serious-Animals A LOAEL, More Serious-Humans I Minimal Risk Level 

r-Rat m-Mouse e-Gerbil @ LOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans I for effects 

h-Rabbit s-Hamster 0NOAEL- Animals A NOAEL- Humans other than 

[%%w a-Sheep g-Guinea Pig Cancer 
mglkglday 

10000 

1000 

100 

10 

1 

0.1 

0.01 

0.001 

0.0001 

c-Cat 

d-Dog

r-Rat 

p-Pig

q-Cow 

Figure 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (Continued) 

Intermediate (15-364 days) 

l s t e m i c 

063r 064r @82r Q84r 086r 

Q75r 

.90h .90h 

037k Q76r 

068r w 

052r 

056r 

@56r 

Q58r Q59r 

@70r @71r 

Q79r 

'*Ir 083r 

080r 

i 

rn 

I 

-n @68r -n 

m 

@41r 042r 

068r Q91p 

041r 042r O38k 040r 

091p 

-Humans f-Ferret n-Mink Cancer Effect Level-Animals V Cancer Effect Level-Humans LD50/LC50

k-Monkey j-Pigeon o-Other LOAEL, More Serious-Animals A LOAEL, More Serious-Humans I Minimal Risk Level 

m-Mouse e-Gerbil Q LOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans : for effects 

h-Rabbit +Hamster 0NOAEL -Animals A NOAEL - Humans A other than 

a-Sheep g-Guinea Pig Cancer 
I 

Figure 3-2. Levels of Significant Exposure to Hexachlorobenzene- Oral (Confinued) 

Intermediate(15-364 days) 

rnglkglday 

10000 

1000 

100 089d 839k 08% 

039k 847r 865r 888; Oglp 

859r 070r 071r 040r 852r 

10 841r 052r 068r 881 r 837k 

858r 059r 891 p 

1 041r 842r 

091p 

0.1 042r 

0.01 

0.001 

0.0001 

T: 

861 r 863r 864r 882r 884 

062r 

089d 037k 062r 

056r 

865r 

852r 

838k 840r 842r 052r 

859r 

08% 

091p 

852r 056r68 870r 871r 

059r 

038k 841r 042r 855r 

091 p 

041 r 

c-Cat -Humans f-Ferret n-Mink Cancer Effect Level-Animals V Cancer Effect Level-Humans LD50/LC50

A LOAEL, More Serious-Humans : Minimal Risk Level .I 

w d-Dog k-Monkey j-Pigeon 0-Other LOAEL, More Serious-Animals r-Rat m-Mouse e-Gerbil 8 LOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans : for effects 

p-Pig h-Rabbit s-Hamster 0NOAEL -Animals A NOAEL - Humans A other than 

Cancer q-Cow a-Sheep g-Guinea Pig 
Figure 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (Continued) 

Intermediate (15-364 days) I

B

0 

c

7J

0 

W 

rn 

z 

N 

rn 

z 

10000 rn 

1000 061 r 064r 

.got1 

037k 100 051r 055r 085r

10 892k 095r 897r 898r 8102r 8104r 5 

091p 093r 0112p I 

F! 

1 

0.1 

0.01 

0.001 

0.0001 

Cancer Effect Level-Animals V Cancer Effect Level-Humans LD50/LC50 c-Cat -Humans f-Ferret n-Mink 

LOAEL, More Serious-Animals A LOAEL, More Serious-Humans I Minimal Risk Level P 

4 &Dog k-Monkey j-Pigeon o-Other LOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans : for effects r-Rat m-Mouse e-Gerbil p-Pig h-Rabbit s-Hamster 0NOAEL - Animals A NOAEL - Humans A other than 

Cancer 
Figure 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (Continued) 

Intermediate (15-364 days) 

mglkglday 

10000 

1000 @114r @115r @I 36r 

100 

@140p 0141k 

w 

0138r 

10 @132k 0138r 

0134k 0140p 

1 @130k 0131k 0135k @137r rn

2 v) 

@146r 

0.1 @128k 0130k @133k @135k 

0.01 9129k 0139r 

0.001 

c-Cat -Humans f-Ferret n-Mink Cancer Effect Level-Animals Cancer Effect Level-Humans LD50/LC50

d-Dog k-Monkey j-Pigeon o-Other LOAEL, More Serious-Animals A LOAEL, More Serious-Humans : Minimal Risk Level 

r-Rat m-Mouse e-Gerbil @ LOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans : foreffects 

p-Pig h-Rabbit s-Hamster 0NOAEL -Animals A NOAEL - Humans A other than 

o-Cow a-Sheeo a-Guinea Pin Cancer 

0.0001 
Figure 3-2. Levels of Significant Exposure to Hexachlorobenzene- Oral (Continued) 

Chronic (2365 days) 

Systemic 

1000 

100 0158d 0158d 

10 

I 0158d 

01 50r 

0158d 0158d 

0153r 

0158d 

0.1 

0.01 

0.001 

0.0001 

IE-5 

1E-6 

1E-7 

1E-8 

c-Cat d-Dog

r-Rat 

p-Pig q-Cow 

-Humans k-Monkey

m-Mouse 

h-Rabbit

a-Sheep 

f-Ferret n-Mink

j-Pigeon o-Other 

e-Gerbil 

s-Hamster

g-Guinea Pig 

Cancer Effect Level-Animals 

LOAEL, More Serious-Animals 

(3LOAEL, Less Serious-Animals 

0NOAEL- Animals 

Cancer Effect Level-Humans 

NOAEL - Humans 

A LOAEL, More Serious-Humans 

A LOAEL, Less Serious-Humans 

01 50r 

w 

I 

rn

5

I 

rn 

n 

n 

rn

2 0, 

LD50/LC50

I Minimal Risk Level 

: for effects 

other than 

Cancer 
Figure 3-2. Levels of Significant Exposure to Hexachlorobenzene - Oral (Confinued) 

Chronic (2365days) 

mglkglday 

1000 

100 

.160m 

10 4168m +166r +167r +I64r 

*169s 4163r *165r 

0161r 0162 

1 

w 

0.1 a159d 

91 62r 0.01 

0.001 

0.0001 "7 Estimated 

1E-5 Upper-Bound 10-4 Human Cancer 

Risk Levels 1 E-6 

1 E-7 *Doses represent the lowest dose tested per study that produced a tumorigenic 

response and do not imply the existence of a threshold for the cancer end point. 

1 E-8 

n-Mink 4 Cancer Effect Level-Animals Cancer Effect Level-Humans LD50/LC50

o-Other LOAEL, More Serious-Animals A LOAEL, More Serious-Humans : Minimal Risk Level 

LOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans A

; for effects 

0NOAEL -Animals NOAEL - Humans other than 

Cancer 
HEXACHLOROBENZENE 78 

3. HEALTH EFFECTS 

lesions, typically occurring together, were found in six different strains of rats and both sexes, at doses as 

low as 3 mg/kg/day in intermediate-duration feeding studies (Goldstein et al. 1978; Kimbrough and 

Linder 1974; Kitchin et al. 1982; Michielsen et al. 1997, 1999, 2001). They were seen at doses as low as 

1 mg/kg/day, with a NOAEL of 0.2 mg/kg/day, in rats that received both pre- and postnatal exposure 

(Vos et al. 1979a, 1983). (The Vos studies are included as developmental toxicity studies in Table 3-2 

and Figure 3-2.) Michielsen et al. (1997) hypothesized an immunomodulated etiology for these lesions, 

but the lesions occurred to a similar extent in five rat strains (Wistar, Lewis, Brown Norway; athymic and 

euthymic WAG/Rij) with very different responses to immunomodulating agents and did not correlate 

with observed immune changes, providing no support for this hypothesis (Michielsen et al. 1997, 1999, 

2001). More severe pulmonary effects have also been observed in rats at higher doses. In addition to 

macrophage accumulation and focal areas of interstitial fibrosis, which they observed in male and female 

rats exposed to $5 mg/kg/day in the diet for 4 months, Kimbrough and Linder (1974) also observed 

extensive intra-alveolar hemorrhage, inflammation, and edema, accompanied by an increase in lung 

weight, in females exposed to 25 or 50 mg/kg/day. Michielsen et al. (2001) observed an adverse effect on 

respiratory function, and airway hyperreponsiveness, as well as granulomatous lung inflammation, in rats 

exposed to 45 mg/kg/day. Only limited pulmonary histopathology data are available for other species, 

and pulmonary lesions were not seen in available studies on monkeys, dogs, or mice (Iatropoulos et al. 

1976; Loose et al. 1977; Sundlof et al. 1981). 

Cardiovascular Effects. No studies were located regarding cardiovascular effects of oral hexachlorobenzene exposure in humans. 

There have been a few reports of cardiovascular lesions in animals exposed to hexachlorobenzene. Gralla 

et al. (1977a) described an arteriopathy affecting multiple organs in dogs treated with 110 mg/kg/day of 

hexachlorobenzene for 1 year. The lesion was characterized by inflammation of small arteries and 

arterioles, with focal proliferative endartitis, fibrinoid necrosis, and thrombosis, and occasionally involved 

fibrosis and inflammation adjacent to the arterioles in the heart and the liver. Although features of the 

lesion suggested a hypersensitivity reaction, an immune etiology was not supported by serum 

electrophoretic data. The arteritis was seen in 4 of 12 beagle dogs treated with 110 mg/kg/day, a dose that 

produced weight loss, mortality, and other frank toxic effects, but was not seen in dogs treated with lower 

doses. It is not known if these effects were produced by a direct effect of hexachlorobenzene or were 

secondary to general poor health of the dogs in this study. However, similar observations in the heart 

were made by Kimbrough and Linder (1974) in rats. These researchers found fibrosis and degeneration 

of muscle fibers in the heart of rats exposed to 25 mg/kg/day or more in the diet for 4 months. 
HEXACHLOROBENZENE 79 

3. HEALTH EFFECTS 

Degenerated tissue was infiltrated by inflammatory cells. Other studies that included pathological 

examination of cardiovascular tissues in dogs, rats, and monkeys did not find treatment-related lesions 

(Goldstein et al. 1978; Iatropoulos et al. 1976; Sundlof et al. 1981). 

Gastrointestinal Effects. No studies were located regarding gastrointestinal effects of oral hexachlorobenzene exposure in humans. 

Gastrointestinal effects have not been commonly reported in animal studies of hexachlorobenzene. Dogs 

given $11 mg/kg/day by capsule for 1 year experienced intermittent episodes of diarrhea (Gralla et al. 

1977a). Necropsy revealed necrotic and inflammatory lesions of the omentum and abdominal serosa, but 

there were apparently no findings in the stomach, small intestines, or large intestines. Pathological 

examination of female pigs exposed to 0.025 or 0.5 mg/kg/day for 212 days throughout mating, gestation, 

and lactation showed gastrointestinal lesions ranging from catarrhal exudation to mild ulceration, but 

microscopic signs of gastritis were also found in some control group animals, suggesting that the 

observed lesions were not an effect of hexachlorobenzene (Hansen et al. 1979). Gastrointestinal lesions 

were not observed in female Rhesus monkeys given oral hexachlorobenzene at doses of up to 

128 mg/kg/day for 60 days (Iatropoulos et al. 1976). 

Hematological Effects. No data were located on the hematological effects of hexachlorobenzene in 

humans. 

Animal data suggest that hexachlorobenzene can produce anemia and leukocytosis. Several studies 

reported decreases in hemoglobin, hematocrit, and/or red blood cell count in rats at doses ranging from 

5 to 32 mg/kg/day in 1–4-month studies (Kimbrough and Linder 1974; Kuiper-Goodman et al. 1977; 

Richter et al. 1981). Female rats were much more sensitive than male rats in these studies (Kimbrough 

and Linder 1974; Kuiper-Goodman et al. 1977). Other studies in rats did not find changes in these 

parameters, but used lower doses (Arnold et al. 1985), a much shorter exposure period (Lecavalier et al. 

1994), or only the less sensitive male rats (Ockner and Schmid 1961). Other reported findings in rats 

consistent with the hypothesis that hexachlorobenzene can produce anemia were increased extramedullary 

hematopoiesis in the spleen, which was seen at doses as low as 5 mg/kg/day with combined pre- and 

postnatal exposure (Vos et al. 1979a, 1979b, 1983), and reduced medullary area in the femur, which was 

found at 10 mg/kg/day and above in a 15-week study (Andrews et al. 1990). Limited data are available 

regarding hematological effects in other species. Anemia was observed in dogs exposed to 

110 mg/kg/day for 1 year (Gralla et al. 1977a), but not in dogs exposed to 100 mg/kg/day for only 
HEXACHLOROBENZENE 80 

3. HEALTH EFFECTS 

3 weeks (Sundlof et al. 1981), in rabbits exposed to 245 mg/kg/day for 12 weeks (De Matteis et al. 1961), 

in pigs exposed to 50 mg/kg/day for 13 weeks (Den Tonkelaar et al. 1978), or in monkeys exposed to 

128 mg/kg/day for 60 days (Knauf and Hobson 1979) or 10 mg/kg/day for 90 days (Foster et al. 1995a). 

Many of the same studies that reported anemia and related findings also reported neutrophilia and/or 

leukocytosis. In rats, the white blood cell increases were found at the same or higher doses than the red 

cell changes ($25 mg/kg/day in 3–16 week feeding studies), and there appeared to be less of a disparity in 

response between males and females (Kimbrough and Linder 1974; Kuiper-Goodman et al. 1977; Vos et 

al. 1979b). In dogs, afebrile neutrophilia was observed at 11 mg/kg/day, while anemia was found only at 

the high dose of 110 mg/kg/day (Gralla et al. 1977a). Negative studies in these species used lower doses 

(Arnold et al. 1985) or a much shorter exposure period (Lecavalier et al. 1994; Sundlof et al. 1981). 

Although values remained within normal ranges, Den Tonkelaar et al. (1978) identified a tendency 

towards leukocytosis and relative neutrophilia in male pigs treated with dietary doses as low as 

0.05 mg/kg/day for 90 days. However, studies in monkeys given doses up to 10 mg/kg/day for 90 days 

(Foster et al. 1995a) or doses up to 128 mg/kg/day for 60 days (Knauf and Hobson 1979) were negative. 

Musculoskeletal Effects. Hexachlorobenzene has been associated with painless arthritis (swelling 

of the joints distinct from rheumatoid arthritis), osteoporosis, and small distinctive hands in patients 

exposed to the chemical from consumption of bread prepared from contaminated grain. Although there 

was severe shortening of digits due to osteoporosis in the bones of the hands (phalangeal, carpal, and 

metacarpal), particularly at the ends, no limitation of movement was reported. Painless arthritic changes 

were also reported in the patients (Cripps et al. 1984; Peters et al. 1982, 1987). 

Effects on both bone and muscle have been reported in animal studies of hexachlorobenzene. Detailed 

studies of bone effects were conducted by Andrews et al. (1989, 1990) in male rats treated by gavage in 

corn oil for up to 15 weeks. These researchers found significant, dose-related increases in femur density 

(osteosclerosis) at doses of 1 mg/kg/day and above, and identified a NOAEL of 0.1 mg/kg/day for this 

effect. Femur length, weight, volume, and cross-sectional area were not consistently altered, indicating 

no effect on the rate of bone growth. The other dose-related changes in bone were an increase in cortical 

area at 1 mg/kg/day and above and a corresponding decrease in medullary area at 10 mg/kg/day and 

above. Other pertinent findings were decreases in serum alkaline phosphatase and increases in serum 

1,25-dihydroxy-vitamin D3 and parathyroid hormone (hyperparathyroidism). The joint findings of 

osteosclerosis, increased cortical and reduced medullary area without a change in the rate of bone growth, 

and decreased serum alkaline phosphatase are consistent with a mechanism involving reduced resorption 
HEXACHLOROBENZENE 81 

3. HEALTH EFFECTS 

of bone. The increases in serum 1,25-dihydroxy-vitamin D3 and parathyroid hormone, both of which are 

involved in calcium regulation and bone resorption, suggest that the hypothesized effect on bone 

resorption is probably secondary to hyperparathyroidism. The decrease in medullary area (bone marrow 

cavity) that ultimately results from these changes may, in turn, contribute to hematological and immune 

effects associated with hexachlorobenzene. 

Studies conducted at higher doses that produced marked interference with heme metabolism found 

substantial accumulation of porphyrins in bone cortex, but not marrow. This was observed in rats treated 

with 100 mg/kg/day in the diet for 56 days (Ockner and Schmid 1961) and rabbits treated with 

245 mg/kg/day in the diet for 84 days (De Matteis et al. 1961). No bone (or liver or kidney) accumulation 

of porphyrins was observed by Andrews et al. (1989), but the highest dose in this study was only 

25 mg/kg/day. 

Skeletal muscle lesions have been reported in animals exposed to hexachlorobenzene, but only with 

repeated exposure to high doses. Rabbits treated with 245 mg/kg/day of hexachlorobenzene for 12 weeks 

were observed to have necrosis, degeneration, and focal calcification in skeletal muscle (De Matteis et al. 

1961). Skeletal muscle lesions were not found in rats fed up to 32 mg/kg/day for 15 weeks (KuiperGoodman et al. 1977). 

Hepatic Effects. The major evidence that oral exposure to hexachlorobenzene by humans can result 

in hepatopathology is derived from studies of an outbreak of porphyria in Turkey attributed to the 

consumption of bread prepared from hexachlorobenzene-contaminated grain from 1955 to 1959 (Cam and 

Nigogosyan 1963; Cripps et al. 1984; Peters et al. 1982, 1987). These adverse hepatic effects were 

mainly characterized by porphyria. The appearance of abnormal levels of porphyrin precursors in the 

urine suggests that hexachlorobenzene disturbed the body's porphyrin metabolism in the liver, which 

caused histopathologic changes in the liver. Uroporphyrin and δ-aminolevulinic acid (d-ALA) synthase 

increased in the urine, and uroporphyrin and coproporphyrin increased in the stool of patients who had 

ingested hexachlorobenzene-contaminated bread (Cripps et al. 1984; Peters et al. 1982, 1987). No 

quantitation of exposure (dose and duration) was presented in any of these clinical reports. However, an 

estimated dose of 0.05–0.2 g/day (0.7–2.9 mg/kg/day for a 70-kg person) is considered to be reliable by 

the original investigators of the Turkey epidemic (Cam and Nigogosyan 1963). 

Studies in animals have confirmed that the liver is an important target organ for hexachlorobenzene 

following ingestion. Hepatic effects associated with oral exposure to hexachlorobenzene in animal 
HEXACHLOROBENZENE 82 

3. HEALTH EFFECTS 

studies include disruption of heme synthesis (culminating in porphyria), induction of microsomal 

enzymes, hepatomegaly, and cellular damage. 

Disruption of hepatic heme synthesis by hexachlorobenzene has been well studied in rats (see below). 

Hexachlorobenzene inhibits the activity of hepatic uroporphyrinogen decarboxylase, an enzyme in the 

heme biosynthesis pathway, leading to build up of heme precursors (porphyrins) in the liver and other 

tissues and their excessive excretion in the urine (porphyria). The activity of other enzymes in the heme 

biosynthesis pathway may also be altered; in particular, an increase in the activity of δ-aminolevulinic 

acid synthestase has been reported in some studies (see Section 3.5, Mechanisms of Action). This pattern 

of effects is very similar to what is seen in human porphyria cutanea tarda. 

Following acute exposure, the lowest dose reported to induce outright porphyria in an animal study was 

25 mg/kg/day in an 8-day study in female rats in which hexachlorobenzene was administered by gavage 

in corn oil; the rats were monitored for urinary and hepatic porphyrins after 35–50 days (Krishnan et al. 

1991). Goldstein et al. (1978) observed a statistically significant increase in hepatic δ-aminolevulinic 

acid synthestase activity in female rats fed $16 mg/kg/day of hexachlorobenzene in the diet for 1 week, 

but there was little or no effect on hepatic porphyrin levels at that time, possibly because of insufficient 

latency time for the effect to develop. (Time course studies have shown that there may be a delay of 

approximately 4 weeks between treatment with hexachlorobenzene and development of porphyria 

[Krishnan et al. 1991; Mylchreest and Charbonneau 1997; Billi de Catabbi et al. 2000a], although there 

may be little or no delay if the hexachlorobenzene is administered in a form that is readily absorbed [e.g., 

predissolved in corn oil] at high doses [Kennedy and Wigfield 1990].) There was no increase in hepatic 

δ-aminolevulinic acid synthestase activity at 5 mg/kg/day in the Goldstein et al. (1978) study. The dose 

level at which hexachlorobenzene will produce porphyria depends on the exposure protocol. When hexachlorobenzene was administered to female rats by gavage for 7 days in an aqueous suspension rather than 

oil, no effect on hepatic uroporphyrinogen decarboxylase (monitored at 7 days) was observed even at 

250 mg/kg/day, and only doses of 500 mg/kg/day or higher were effective (Kleiman de Pisarev et al. 

1990), reflecting the fact that hexachlorobenzene administered in water is only minimally absorbed from 

the gastrointestinal tract, and also perhaps, the short latency time. Even at 1,000 mg/kg/day in this study, 

uroporphyrinogen decarboxylase activity was decreased only 25%, while δ-aminolevulinic acid 

synthestase activity and liver porphyrin levels were unchanged from controls. Billi de Catabbi et al. 

(2000a) observed that acute (5-day) exposure above a threshold (1 g/kg) caused persistent porphyria 

(lasting at least as long as the 20-week observation period). 
HEXACHLOROBENZENE 83 

3. HEALTH EFFECTS 

Hexachlorobenzene doses as low as 5–12 mg/kg/day have been reported to produce porphyrinogenic 

effects, such as increased liver weight, inhibition of hepatic uroporphyrinogen decarboxylase, 

accumulation of porphyrins in liver, excretion of porphyrins in urine, and increased hepatic 

δ-aminolevulinic acid synthestase activity, in female rats exposed for intermediate durations (Alvarez et 

al. 2000; Den Besten et al. 1993; Goldstein et al. 1978; Kennedy and Wigfield 1990; Kuiper-Goodman et 

al. 1977; Smith et al. 1979, 1985; Sweeney et al. 1986). One of these studies, Goldstein et al. (1978), 

identified a NOAEL of 4 mg/kg/day for increases in liver porphyrins and δ-aminolevulinic acid 

synthestase activity after a 4-month dietary exposure. In male rats, there was little or no evidence of 

porphyria at doses up to 25 mg/kg/day (Andrews et al. 1989; Kuiper-Goodman et al. 1977; Smith et al. 

1985), but mild changes were noted at 32–50 mg/kg/day (Krishnan et al. 1991; Kuiper-Goodman et al. 

1977) and severe porphyria was observed at 100 mg/kg/day (Ockner and Schmid 1961). The reduced 

sensitivity of male rats in comparison to females may be related to differences in metabolism of hexachlorobenzene between the sexes in this species, particularly with regard to glutathione conjugation 

(D’Amour and Charbonneau 1992; Richter et al. 1981; Rizzardini and Smith 1982), which have been 

linked to the presence of estradiol in the females (Legault et al. 1997). Among female rats, strain-related 

differences in sensitivity have also been reported (Billi de Catabbi et al. 2000a; Michielsen et al. 1997), 

possibly related, at least in one case, to differences in nonheme iron content of the liver between strains 

(Smith et al. 1979). In other species, there was no evidence of porphyria in female monkeys treated with 

up to 10 mg/kg/day by capsule for 3 months (Jarrell et al. 1993) or in male pigs exposed to up to 

50 mg/kg/day in the diet for 3 months (Den Tonkelaar et al. 1978); these results appear to reflect species 

differences, but may be influenced by the lack of oil vehicle to enhance absorption. 

In studies of chronic exposure duration, hexachlorobenzene doses of 7–10 mg/kg/day in the feed 

produced complete inhibition of uroporphyrinogen decarboxylase and high levels of porphyrins in the 

liver and urine in both male and female rats (Smith and Cabral 1980; Smith et al. 1985, 1993). Although 

uroporphyrinogen decarboxylase activity was completely inhibited in both sexes, liver accumulation of 

porphyrins was 5-fold higher in females than in males (Smith et al. 1985). Chronic rat studies that 

employed lower dose levels did not monitor porphyrin levels (Arnold et al. 1985; Mollenhauser et al. 

1975). Male mice exposed to slightly higher dietary doses of hexachlorobenzene (13 mg/kg/day) showed 

only a modest transitory increase in hepatic porphyrin levels after 6 months of treatment, which was not 

found at subsequent sacrifices at 12 and 18 months (Smith et al. 1989). There was no evidence of 

porphyrin accumulation in the liver or other tissues of dogs treated with hexachlorobenzene doses as high 

as 110 mg/kg/day for 1 year (Gralla et al. 1977a, 1977b). 
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As in human porphyria cutanea tarda, porphyria in rats produced by hexachlorobenzene typically occurs 

along with other effects on the liver, such as induction of microsomal enzymes, increased liver weight, 

hepatocellular hypertrophy, cytoplasmic vacuolation, fatty degeneration, and biliary hyperplasia (Cuomo 

et al. 1991; Den Besten et al. 1993; Kimbrough and Linder 1974; Koss et al. 1978; Kuiper-Goodman et 

al. 1977; Michielsen et al. 1997; Smith et al. 1985, 1993; Sweeney et al. 1986; Vos et al. 1979b). The 

relationship between porphyria and these other hepatic effects is uncertain. In some instances, effects on 

liver weight, enzymes, and/or histopathology occurred in rats at lower doses than porphyria or in the 

absence of porphyria (e.g., Arnold et al. 1985; Gustafson et al. 2000; Kishima et al. 2000; Mehendale et 

al. 1975; Mollenhauer et al. 1975; Michielsen et al. 2000). Liver lesions have also been observed in 

species, such as monkeys, dogs, and pigs, where there is no evidence of a porphyrinogenic effect (Den 

Tonkelaar et al. 1978; Gralla et al. 1977a; Iatropoulos et al. 1976; Jarrell et al. 1993). Kishima et al. 

(2000) reported that the hepatotoxicity of hexachlorobenzene was increased by an energy-restricted diet. 

The most sensitive hepatic end points following acute, intermediate, and chronic exposure, respectively, 

were increased liver weight and induction of microsomal enzymes in male rats exposed to 10 mg/kg/day 

for 6 days (Mehendale et al. 1975), hepatocellular hypertrophy in male pigs exposed to 0.5 mg/kg/day for 

90 days (Den Tonkelaar et al. 1978), and peribiliary lymphocytosis and fibrosis in F1 adult male rats 

exposed to 0.016 mg/kg/day in utero and throughout their lifetime in a 2-generation study (Arnold et al. 

1985; listed as a developmental effect in Table 3-2 and the basis for the chronic oral MRL of 

0.00005 mg/kg/day, as described in the footnote to Table 3-2 and in Appendix A). 

Renal Effects. No studies were found regarding renal effects in humans following oral exposure to 

hexachlorobenzene. 

Animal studies have demonstrated that the kidney is a target for hexachlorobenzene. Renal effects that 

have been widely reported in animal studies are increased kidney weight, accumulation of porphyrins in 

association with disruption of heme metabolism (as in the liver), and direct and indirect evidence of renal 

tissue damage. Increases in kidney weight have been observed in many studies, primarily those involving 

$7 weeks of exposure (Andrews et al. 1989; Bouthillier et al. 1991; Den Besten et al. 1993; Den 

Tonkelaar et al. 1978; Kimbrough and Linder 1974; Koss et al. 1978; Kuiper-Goodman et al. 1977; Smith 

et al. 1985). Most studies shorter than 7 weeks in duration did not find increases in kidney weight, even 

using doses as high as 100 mg/kg/day (Andrews et al. 1988; Richter et al. 1981; Sundlof et al. 1981; Vos 

et al. 1979b). This includes interim sacrifices in longer duration studies that did eventually show 

increases in kidney weight (Andrews et al. 1989). Multiple-dose feeding studies of 12–16 weeks 

identified LOAEL values of 19–32 mg/kg/day and NOAEL values of 5–9.5 mg/kg/day for increased 
HEXACHLOROBENZENE 85 

3. HEALTH EFFECTS 

kidney weight in male and female rats (Den Besten et al. 1993; Kimbrough and Linder 1974; KuiperGoodman et al. 1977). Lower critical values were obtained when exposure was by gavage in oil rather 

than in the diet (Andrews et al. 1989, 1990), when pigs were tested rather than rats (Den Tonkelaar et al. 

1978), and when exposure duration was extended over 1 year (Smith et al. 1985). The lowest LOAEL 

and NOAEL for increased kidney weight in any study were 1 and 0.1 mg/kg/day, respectively (Andrews 

et al. 1989). It has been proposed that increased kidney weight in animals exposed to hexachlorobenzene 

may result from induction of renal microsomal enzymes (Bouthillier et al. 1991). Renal pathology 

(described below) may have been a contributing factor in some studies. 

Among the available studies, the most sensitive indication of renal pathology due to hexachlorobenzene 

was increased urine enzyme levels. Male rats treated with $1 mg/kg/day of hexachlorobenzene by 

gavage in oil for 15 weeks had increased concentrations of alkaline phosphatase and/or lactate 

dehydrogenase in the urine, an effect that was not found at 0.1 mg/kg/day (Andrews et al. 1989, 1990). 

These effect levels correspond with those for changes in kidney weight in the same studies. Increased 

urinary levels of these enzymes suggest the occurrence of either glomerular damage allowing the enzymes 

to leak into the urine from the serum or tubular cell damage where the enzymes are released directly from 

the damaged cells into the urine. There was an apparent increase in calcium excretion in a higher dose 

group that could be interpreted to indicate impaired reabsorption of calcium by the distal tubules, which 

would support the hypothesis that damaged tubular cells were responsible for the observed enzymuria. 

Studies that included histopathological examination of the kidneys provided more direct evidence of 

damage to renal tubule cells. Bouthillier et al. (1991) observed degenerative and regenerative foci of 

epithelial cells in the proximal tubules and accumulation of protein droplets in proximal tubular cells in 

male rats treated with 50 mg/kg/day for 7 weeks or 100 mg/kg/day for 2 weeks. These lesions were 

accompanied by glucosuria, proteinuria, and an 11-fold increase in α2μ-globulin levels. The nature of the 

effects in males and near-absence of effects in females (just glucosuria and an increase in urinary 

γ-glutamyl transpeptidase were found) led the researchers to suggest that hexachlorobenzene produces a 

male rat specific protein droplet nephropathy, as is seen with some other chlorinated benzenes. Support 

for a nephrotoxic effect in rats is provided by Smith et al. (1985), who found mild to severe nephrosis in 

12/12 surviving male rats treated with 10 mg/kg/day for 90 weeks, but 0/10 surviving male controls and 

only 1/10 surviving treated females. Similarly, Arnold et al. (1985) observed chronic nephrosis in male, 

but not female, rats with combined pre- and postnatal lifetime exposure to 2 mg/kg/day of hexachlorobenzene. 
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Renal lesions, however, have been reported in female rats treated with hexachlorobenzene. Basophilic 

renal tubules and protein casts were seen in female rats exposed to 19 mg/kg/day for 13 weeks (Den 

Besten et al. 1993). Renal accumulation of porphyrins has been reported in female rats exposed to doses 

as low as 5 mg/kg/day for 56 days (see below). Renal effects have also been noted in other animal 

species. Female monkeys treated with $8 mg/kg/day for 60 days developed vacuolization of proximal 

tubules, and at 128 mg/kg/day, thickening of the basement membranes, glomerular hyperemia, and 

increased BUN were also found (Iatropoulos et al. 1976; Knauf and Hobson 1979). Male pigs exposed to 

50 mg/kg/day in the diet all died prior to scheduled sacrifice and were found upon necropsy to have 

degeneration of the proximal tubules and loop of Henle (Den Tonkelaar et al. 1978). Blood ammonia 

levels were increased 3-fold in female guinea pigs treated with 200 mg/kg/day (De Matteis et al. 1961). 

There is also a brief paper by Ertürk et al. (1986), lacking detail regarding experimental methods or 

results, that reports marked renal lesions, including severe hyperemia, necrotic tubular degeneration, 

hemorrhage, and nephritis in male and female rats, mice, and hamsters given 90-day dietary exposure to 

10–33 mg/kg/day (although the effects were most severe in male rats, they were seen in all groups). The 

occurrence of renal effects in female rats and in other species of laboratory animal shows that the 

nephrotoxic effects of hexachlorobenzene are not limited to the α2μ-globulin nephropathy demonstrated 

by Bouthillier et al. (1991), and suggests that this chemical may produce renal effects by multiple 

mechanisms. 

Renal accumulation of porphyrins has been related to disruption of heme metabolism and lipid 

peroxidation. Female rats treated with 1,000 mg/kg/day of hexachlorobenzene by gavage in aqueous 

Tween 20 had a significant decrease in uroporphyrinogen decarboxylase (URO-D) activity in the renal 

cortex after 3 weeks of exposure and a subsequent increase in porphyrin levels in the renal cortex, but not 

the renal medulla or papilla, after 4 weeks of exposure (Fernandez-Tome et al. 2000). Lipid peroxidation, 

indicated by measurement of conjugated dienes and malondialdehyde (MDA), was significantly increased 

throughout most of the exposure period in the renal cortex, but was not increased at all in the renal 

medulla or papilla. Based on these findings, the researchers suggested that disruption of heme 

metabolism and accumulation of porphyrins in the renal cortex are secondary to lipid peroxidation 

produced by hexachlorobenzene in this tissue. Distribution of porphyria in kidney is consistent with fact 

that enzymes of heme metabolism are localized in the renal cortex; the occurrence of lipid peroxidation in 

the renal cortex is consistent with its relative susceptibility to oxidative stress, compared to the papilla or 

medulla. The use of such a high dose in this study reflects the fact that hexachlorobenzene is not well 

absorbed from water. Renal accumulation of porphyrins was observed at much lower doses (as low as 

5 mg/kg/day) in feeding studies in female rats (Kennedy and Wigfield 1990; Smith et al. 1985), 
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3. HEALTH EFFECTS 

particularly when hexachlorobenzene was added to the diet in an oil vehicle (Kennedy and Wigfield 

1990). No experimental evidence was found for renal accumulation of porphyrins in male rats treated 

with up to 25 mg/kg/day for 15 weeks in the diet or by gavage in oil (Andrews et al. 1989; Smith et al. 

1985). 

Despite the numerous data supporting an effect of hexachlorobenzene on the kidney, it should be noted 

that several well-conducted investigations of kidney histopathology failed to find any treatment-related 

lesions in either male or female rats, even with high exposures (up to 50 mg/kg/day for 4 months) that, 

based on the database as a whole, would have been expected to produce tissue damage (Kimbrough and 

Linder 1974; Kuiper-Goodman et al. 1977; Michielsen et al. 1997; Vos et al. 1979b). 

Endocrine Effects. Human data suggest that hexachlorobenzene adversely affects the endocrine 

system; specifically, the thyroid is a target organ. Two follow-up studies conducted 25 (Peters et al. 

1982) and 20–30 years (Cripps et al. 1984) after patients (n=161–225) were exposed as children to bread 

contaminated with hexachlorobenzene in Southeast Turkey detected thyromegaly in 60% of women and 

25% of men (Cripps et al. 1984; Peters et al. 1982). The background incidence for this area was 5%. 

Additionally, hirsutism and small stature were observed in 47 and 44%, respectively, of the study 

population. However, a study of serum hormone and organochlorine levels in a cohort of 110 Swedish or 

Latvian men who consumed Baltic sea fish (Hagmar et al. 2001) found no age-adjusted correlation 

between hexachlorobenzene levels and any of the measured serum hormones (follicle-stimulating 

hormone, luteinizing hormone, prolactin, plasma thyrotropin, free and total T3, free and total T4, and free 

testosterone). The authors considered the fish to be a diet high in persistent organohalogens, but did not 

quantitate intake. 

Animal studies have demonstrated that hexachlorobenzene has multiple endocrine effects; the most 

striking are the induction of hypothyroidism and hyperparathyroidism in rats. Limited evidence suggests 

that hexachlorobenzene also affects serum retinoid levels, the adrenal gland and serum levels of 

corticosterone and cortisol (see below). Moreover, studies have shown that hexachlorobenzene affects 

serum levels of estrogen and progesterone (discussed in Section 3.2.2.5, Reproductive Effects). 

Multiple studies have demonstrated that serum T4 levels decrease rapidly in rats following gavage 

treatment with hexachlorobenzene (Den Besten et al. 1993; Den Tonkelaar et al. 1978; Foster et al. 1993; 

Kleiman de Pisarev et al. 1989, 1990, 1995; Smith et al. 1986b; Sopena de Kracoff et al. 1994; van Raaij 

et al. 1993a, 1993b). Effects on serum TSH levels (both increases and decreases) are delayed and appear 
HEXACHLOROBENZENE 88 

3. HEALTH EFFECTS 

secondary to decreases in T4. The most sensitive acute study observed statistically significant decreases 

in T4 levels in female rats at doses as low as 50 mg/kg/day for 5 days (Foster et al. 1993). A time-course 

in female Wistar rats gavaged with 1,000 mg/kg/day of hexachlorobenzene in corn oil found that serum 

T4 levels rapidly decreased, reaching a steady-state after 8 days, approximately 75% below controls. In 

contrast, serum TSH levels reached a steady state after 30 days, at 80% below controls (Sopena de 

Kracoff et al. 1994). Similarly, experiments in female Wistar rats gavaged with 1,000 mg/kg/day for at 

least 8 weeks observed significantly decreased serum T4 and protein-bound iodine and elevated TSH 

levels and thyroid weight (Kleiman de Pisarev et al. 1989, 1990, 1995). 

The most sensitive intermediate-duration studies were conducted in male Syrian hamsters (Smith et al. 

1986b, 1987), which responded differently than rats to hexachlorobenzene. In hamsters exposed to 

10 mg/kg/day in feed for 28 weeks (the lowest effective intermediate-duration dose identified) or to a 

higher dose for a shorter time (50 mg/kg/day or more for 6 weeks), thyroid gland weights were 

significantly increased ($2.5-fold), serum T3 levels were decreased, and sodium iodide uptake was 

increased (~3-fold), while serum T4 levels were unchanged. Thyroid weight increased significantly, and 

correlated with histopathological observations of large and irregularly shaped follicles in the thyroid. 

Notably, two other studies did not observe treatment-related thyroid histopathology at effective doses; 

Den Besten et al. (1993) observed decreased T3 and T4 levels in female Wistar rats exposed to hexachlorobenzene in the feed at doses up to 19.0 mg/kg/day for 13 weeks, and Den Tonkelaar et al. (1978) 

detected significantly increased thyroid weight in male Landvarken pigs fed 5 mg/kg/day of hexachlorobenzene for 12 weeks. 

Increased hepatic metabolism and hepatic excretion (into bile) appear to be important to the mode-ofaction for the thyroid effects of hexachlorobenzene. Kleiman de Pisarev et al. (1989, 1990, 1995) 

observed that gavage administration of 1,000 mg/kg/day of hexachlorobenzene in corn oil to female 

Wistar rats for 28–30 days not only significantly decreased serum T4 levels, but also increased both the 

metabolism (deiodination) and fecal excretion of T4. In male Wistar (WAG/MBL) rats orally dosed with 

120 mg/kg/day of hexachlorobenzene 3 times/week for 4 weeks, levels of T4 glucuronide increased but 

levels of serum T4 and nonconjugated T4 decreased (van Raaij et al. 1993b). Hepatic T4 UDP-glucuronyltransferase (UDPGT) activity was increased while T4 iodothyronine deiodinase activity was 

decreased. Bile flow and T4 excretion were increased. However, serum T3 levels were unaffected. 

Taken together, these data indicate that hexachlorobenzene induces liver activity that results in decreased 

serum levels of T4. Treatment of female Sprague-Dawley rats with 50 mg/kg/day of hexachlorobenzene 

by gavage for 5 days significantly decreased serum T4 levels and free thyroxine index without affecting 
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3. HEALTH EFFECTS 

T3 uptake (Foster et al. 1993). However, following gonadotropin pretreatment to induce superovulation, 

3-day treatment with 50 mg/kg/day of hexachlorobenzene significantly decreased both serum T4 levels 

and serum T3 uptake. The authors concluded that the acute induction of hypothyroidism was augmented 

by ovulation (Foster et al. 1993). Gavage treatment of inbred male Wistar (WAG/RIJ) rats with 

27 mg/kg/day pentachlorophenol was more effective than 780 mg/kg/day of hexachlorobenzene at 

reducing free and total T4 levels (van Raaij et al. 1993a). The authors concluded that metabolites of 

hexachlorobenzene, rather than hexachlorobenzene itself, may be involved in the reduction of serum 

thyroid hormones. 

Andrews et al. (1988, 1989, 1990) have conducted detailed studies into hexachlorobenzene-induced 

hyperparathyroidism and osteoporosis in male Fischer 344 rats. Gavage treatment of rats with doses as 

low as 1.0 mg/kg for 5 weeks, 5 days/week, significantly increased serum levels of vitamin D3 and 

urinary levels of phosphorous. Treatment with at least 10 mg/kg significantly increased serum levels of 

PTH, increased urinary levels of calcium, and decreased serum levels of alkaline phosphatase (an enzyme 

important in bone mineralization). The combination of high PTH and vitamin D3 is expected to cause 

calcium resorption from bone and calcium conservation by the kidneys, and this is consistent with the 

adverse skeletal effects seen by authors at 1.0 mg/kg (see Section 3.2.2.2, Musculoskeletal Effects). 

Animal studies have shown that hexachlorobenzene affects the adrenal gland (weight, histopathology, and 

hormone levels). Adrenal gland weight was significantly increased in male and female Sherman rats fed 

at least 25 mg/kg/day of hexachlorobenzene for 4 months (Kimbrough and Linder 1974). Adrenal gland 

weight was also significantly increased in Wistar rat pups exposed to at least 2.5 mg/kg/day of hexachlorobenzene (starting early in pregnancy and continuing through gestation and lactation, and an 

additional 2 weeks after weaning) (Vos et al. 1979a). In female Wistar rats that were fed 9.5 or 

19.0 mg/kg/day of hexachlorobenzene for 13 weeks (Den Besten et al. 1993), adrenal gland weight (69%) 

was significantly increased in the high-dose group. Histopathology (reported only for the high-dose 

group) observed adrenal cortex hypertrophy, hyperplasia, occasional hemorrhaging, cortical cell 

vacuolation, and inflammatory cell infiltrates. In female ovariectomized Sprague-Dawley rats gavaged 

for 30 days with 1, 10, or 100 mg/kg/day of hexachlorobenzene, corticosterone was decreased at all 

doses, but serum cortisol levels were only decreased by 100 mg/kg/day (Foster et al. 1995a). Moreover, 

no effect was seen on serum progesterone and aldosterone levels or adrenal weight. Koss et al. (1978) 

detected statistically significant increases in relative adrenal weight (up to 43%) in female Wistar rats that 

were treated with 50 mg/kg hexachlorobenzene every other day by gavage for 9–15 weeks; this effect had 

reversed after a 38-week posttreatment period. In another experiment, single adult female monkeys 
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3. HEALTH EFFECTS 

received 0, 8, 32, or 64 mg/kg and two monkeys received 128 mg/kg of hexachlorobenzene by gavage in 

methylcellulose daily for 60 days (Iatropoulos et al. 1976). One of the monkeys given 128 mg/kg 

exhibited moderate adrenal medullary hyperplasia and the monkey given 64 mg/kg/ exhibited slight 

hyperplasia of the adrenal zona fasciculata; however, these findings are not conclusive due to the small 

numbers of animals used. 

Limited animal evidence suggests that hexachlorobenzene affects retinoid levels. In blood samples taken 

from 101 polar bears from Svalbard, Norway, statistically significant correlations were observed between 

higher blood levels of hexachlorobenzene and both lower levels of retinol and a lower ratio of total 

thyroxine (T4) to free triiodothyronine (T3) (Skaare et al. 2001). In female Wistar rats fed 19 (but not 

9.5) mg/kg/day of hexachlorobenzene for 13 weeks (Den Besten et al. 1993), significant increases were 

seen in adrenal gland weight (69%) and both liver retinol and retinyl palmitate levels. Plasma retinol 

levels were not affected at 1 week, but were significantly increased at 13 weeks by the high dose. 

Dermal Effects. Studies of humans exposed to hexachlorobenzene in bread prepared from 

contaminated grain in Turkey demonstrated that hexachlorobenzene can produce skin lesions following 

oral exposure. It is well known that ingestion of hexachlorobenzene can produce porphyria (see 

Section 3.2.2.2, Hepatic Effects). The porphyrias are a class of inherited and acquired diseases caused by 

enzymatic defects in heme biosynthesis, leading to the generation of porphyrins. One of the most serious 

symptoms of porphyria is photosensitivity; porphyrins accumulated in the skin absorb radiation 

(maximally at 400–410 nm) and then generate reactive oxygen species, causing tissue damage (Lim and 

Cohen 1999; Meola and Lim 1993; Sandberg et al. 1982). Skin lesions occur most commonly on areas 

exposed to sunlight, such as the hands and face. Porphyria cutanea tarda, a specific type of 

vesiculobullous porphyria, was widespread in southeast Anatolia, Turkey in the late 1950s 

(approximately 1955–1959). The disease was traced to ingestion of bread made from seed grain that had 

been treated with hexachlorobenzene as a fungicide (Cam and Nigogosyan 1963). The ingested dose of 

hexachlorobenzene by exposed persons was estimated to be in the range of 0.05–0.2 g/day (0.7–2.9 mg/ 

kg/day for a 70-kg person) during the episode (Cam and Nigogosyan 1963). Symptoms of what was 

called kara yara or “black sore” appeared after approximately 6 months of exposure (Gocmen et al. 

1989). The disease was observed most frequently in children between the ages of 6 and 15 years, 

although some younger children and adults were also affected (Cam and Nigogosyan 1963; Dogramaci 

1964). The initial lesions resembled comedones (blackhead acne) and milia (small whitish epidermal 

cysts caused by hair follicle and sweat gland obstruction) with photosensitivity and the development of 

erythema on sun-exposed areas; moreover, the skin was sensitive to minor trauma. These lesions 
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3. HEALTH EFFECTS 

progressed to include large bullous lesions that ulcerated and healed leaving severe mutilating scars, 

hyperpigmentation that was most prominent on exposed areas but usually affected the entire skin, and 

hypertrichosis (hirsutism) that occurred principally on the forehead, cheeks, arms, and legs but 

occasionally involved the whole body (Gocmen et al. 1989). Infants who had been breast fed by mothers 

who had ingested the contaminated bread displayed a condition known as pembe yara or "pink sore" 

because of the associated skin lesions (annular erythema) (Peters et al. 1966, 1982). The medical history 

of people who were exposed to hexachlorobenzene in the Turkish poisoning episode was followed for up 

to 30 years (Cripps et al. 1984; Gocmen et al. 1989; Peters et al. 1982, 1987). A total of 252 patients who 

had become porphyric during the Turkish epidemic were studied over a 10-year period from 1977 to 1987 

(Gocmen et al. 1989). Subjects had an average age of 7.6 years at onset of symptoms and an average age 

at follow-up of 35.7 years. Thirty years after onset, clinical findings persisted in most subjects, including 

severe scarring (83.7%), hyperpigmentation (65%), hypertrichosis (44.8%), pinched face appearance with 

perioral scarring (40.1%), and fragile skin (33.7%). Dermal lesions were more prominent in sun-exposed 

areas of the skin. 

Studies into the mode of action of porphyria (unrelated to hexachlorobenzene-exposure) have suggested 

that the dermal toxicity of porphyrins is exacerbated by involvement of the immune system; following 

irradiation, porphyrins may activate the complement system and stimulate mast cells and neutrophils to 

damage nearby tissues (Lim and Cohen 1999; Meola and Lim 1993). Additionally, an increased risk of 

porphyria cutanea tarda has been associated with human immunodeficiency virus (HIV) infection 

(Drobacheff et al. 1998; Egger et al. 2002). 

Although several animal studies have demonstrated that oral exposure to hexachlorobenzene (for at least 

4 weeks) results in dermal lesions and causes immunological effects with dermal lesions, none established 

a casual relationship (Koss et al. 1978; Michielsen et al. 1997, 2000; Schielen et al. 1995a; Torinuki et al. 

1981). Additionally, it is unclear whether porphyrin-induced phototoxicity occurs in rats; the 

combination of hexachlorobenzene and sunlight exposure induced dermal lesions in rats similar to those 

reported in people (Torinuki et al. 1981). However, hexachlorobenzene-treated rats have exhibited skin 

lesions without detectable dermal porphyrin accumulation (assayed with fluorescence microscopy) 

(Michielesen et al. 1997). Dermal lesions in rats following hexachlorobenzene exposure have been seen 

most frequently on the head, neck, and shoulders (similar to humans) although the rats’ exposure to 

sunlight was presumably limited by experimental design and body fur (Koss et al. 1978; Michielsen et al. 

1997, 2000). These data seem to suggest that additional modes-of-action are important. 
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3. HEALTH EFFECTS 

Dermal effects were noted in a study of female Wistar rats that were gavaged with 50 mg/kg hexachlorobenzene in olive oil every other day for 9–15 weeks, followed by a 38-week observation period (Koss et 

al. 1978). The fur had a roughened appearance during treatment, and round, ulcerous lesions on the head, 

ears, throat, and shoulders, with diameters of 2–20 mm, were observed in 10% of treated animals after 

4 weeks and 50% of animals after 9 weeks. These lesions resolved 12–16 weeks after discontinuation of 

treatments. At 15 weeks only, spleen weight was significantly increased. A subsequent study, using 

lower doses and more sensitive end points, observed strain specificity in rats fed hexachlorobenzene for 

4 weeks (Michielsen et al. 1997). Brown Norway and Lewis rats received 18 or 53 mg/kg/day, while 

Wistar rats were fed 50 or 97 mg/kg/day. Increased incidence of skin lesions were statistically significant 

for all treatment groups compared to controls. In Brown Norway rats, skin lesions were very severe, and 

their incidence was correlated with signs of immunomodulation (increased IgG, IgE, and IgM levels, 

spleen weight, and lung inflammation). In Lewis rats, skin lesions were moderate, and correlated less 

strongly with evidence of immune effects. In Wistar rats, lesions were observed only at the high dose, 

and were considered negligible; immune effects were also minor. Grossly, the lesions were found in the 

head and neck region and ranged from redness only to large exudating sores with crusts. Histopathology 

of both lesional and non-lesional skin observed epidermal hyperplasia, deep dermal venules with 

activated endothelial cells, and deep inflammatory cell infiltrates. A subsequent experiment (Michielson 

et al. 2000) verified that doses as low as 58 mg/kg/day in the feed of female Brown Norway rats for 

4 weeks induced skin lesions in the head and neck areas, the severity of which increased with time. 

Likewise, a study in female Wistar rats fed 15 or 30 mg/kg/day of hexachlorobenzene for 13 weeks 

observed statistically significant increases in the incidence of wounds appearing by weeks 5–7 on the 

face, neck, shoulders, and behind the ears of treated animals (Schielen et al. 1995a). The incidence (but 

not severity) increased with increasing dose. Treatment also induced porphyria and increased serum 

levels of IgM, IgA, and autoantigen-specific IgM. Torinuki et al. (1981) treated rats for 2 months with 

hexachlorobenzene with repeated exposure to sunlight. In addition to porphyria, gross pathology of the 

skin observed erythema, erosion, crusting, skin thickening, and scarring. Histopathology found 

acanthosis (abnormal dermal thickening), vacuolization of malpighian cells, subepidermal vesicles, blood 

vessel dilation, and perivascular cell infiltration of lymphocytes, histiocytes, and mast cells. In addition 

to lesions, hexachlorobenzene has induced dermal effects that are not clearly toxic. In female Wistar rats, 

skin cytochrome P450-dependent 7-ethoxyresorufin-O-deethylase was induced after 60 or 70 days (but 

not 10 days) of feeding with 50 mg/kg/day of hexachlorobenzene (Goerz et al. 1978, 1994). No dermal 

lesions were observed in female Agus Wistar rats fed diets containing 5 mg/kg/day of hexachlorobenzene 

for 90 weeks, although treated animals “possessed less hair than controls” (Smith and Cabral 1980). 
HEXACHLOROBENZENE 93 

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Ocular Effects. No studies were found regarding adverse ocular effects in humans following oral 

exposure to hexachlorobenzene. 

Only one relevant animal study was identified. In adult female Rhesus monkeys treated by gavage with 

hexachlorobenzene in methylcellulose at doses up to 32 mg/kg/day for 60 days, gross pathology and 

histopathology of the eyes did not detect any adverse effects (Iatropoulos et al. 1976). 

Body Weight Effects. No studies were located regarding body weight effects in humans following 

oral exposure to hexachlorobenzene. 

Decreases in body weight following oral exposure of animals to hexachlorobenzene were not observed 

following acute exposure and were observed in intermediate- and chronic-duration experiments only in 

the presence of other adverse effects such as mortality, clinical evidence of weakness or lethargy, 

increased incidences of tumors, or liver toxicity (see below and Section 3.3.2). The available data suggest 

that body weight loss may be secondary to organ-specific toxicity. 

No acute studies have reported changes in body weight. Body weight was not affected in rats dosed once 

by gavage with 400 or 600 mg/kg of hexachlorobenzene in corn oil (Lecavalier et al. 1994) or in male rats 

gavaged with doses up to 221 mg/kg/day for 5 consecutive days (Simon et al. 1979). 

Similarly, many intermediate-duration experiments did not detect significant changes in body weight in 

Wistar rats fed up to 1,000 mg/kg/day of hexachlorobenzene for 4 weeks (Alvarez et al. 2000; Kleiman de 

Pisarev et al. 1995; Michielsen et al. 1997; Schielen et al. 1993, 1995b; Vos et al. 1979b), in Brown 

Norway rats and Lewis rats fed up to 45 mg/kg/day for 4 weeks (Michielsen et al. 1997, 2000), in 

ovariectomized adult female Sprague-Dawley rats gavaged with doses up to 100 mg/kg/day of hexachlorobenzene in corn oil for 30 days (Foster et al. 1995b), in female CD rats fed up to 36 mg/kg/day for 

4 months (Goldstein et al. 1978), in female Wistar rats fed 5 mg/kg/day of hexachlorobenzene in arachis 

oil for 75 weeks (Smith and Cabral 1980), or in male BALB/c mice fed 22 mg/kg/day for 6 weeks (Loose 

et al. 1977). Intermediate-duration experiments that did report statistically significant weight loss in rats 

(Kuiper-Goodman et al. 1977; Smith and Cabral 1980; Smith et al. 1985) or mice (Cabral et al. 1979; 

Shirai et al. 1978) usually considered it to be slight-to-moderate. The most rapid onset of weight loss was 

a slight decrease at doses as low as 22 mg/kg/day in male Wistar rats for 6 weeks, in the presence of 

increased liver, spleen, and lymph node weight and decreased natural killer cell activity (Van Loveren et 

al. 1990). Although one study reported statistically significant decreases in body weight in male 
HEXACHLOROBENZENE 94 

3. HEALTH EFFECTS 

Fisher 344 rats at doses as low as 0.1 mg/kg of hexachlorobenzene by gavage in corn oil 5 days/week for 

5 weeks (Andrews et al. 1988), subsequent studies using the same protocols in 15-week experiments saw 

no effects on body weight (Andrews et al. 1989, 1990). Oral intermediate-dosing studies that observed 

hexachlorobenzene-induced mortality also found significant weight loss. In a 13-week feeding study, 

four of nine female Wistar rats receiving the high dose of 19 mg/kg/day of hexachlorobenzene in corn oil 

for 13 weeks were euthanized during the study due to severe weight loss and distress, while no body 

weight effects were seen in surviving or lower dose animals (Den Besten et al. 1993). Male Syrian 

golden hamsters fed 16 (but not lower doses up to 8) mg/kg/day of hexachlorobenzene in the diet for life, 

males exhibited “marked weight reduction” and significantly increased mortality (Cabral et al. 1977). 

However, similar mortality was observed in treated female hamsters without corresponding weight loss. 

Similar findings have also been demonstrated in dogs, monkeys, and pigs. No treatment-related effects 

on body weight were observed in adult female beagle dogs that received hexachlorobenzene in corn oil 

orally in capsules for 3 weeks at doses up to 100 mg/kg/day (Sundlof et al. 1981) or in male and female 

beagle dogs fed capsules of hexachlorobenzene at doses up to 110 mg/kg/day for up to 12 months (Gralla 

et al. 1977a). Similarly, experiments in which female Cynomolgus monkeys were dosed with hexachlorobenzene in gelatin capsules at doses up to 10 mg/kg/day for 90 days, did not detect changes in body 

weight (Foster et al. 1992a, 1995a; Jarrell et al. 1993). In contrast, unspecified weight loss was reported 

beginning at 4 weeks in adult female Rhesus monkeys given hexachlorobenzene by daily gavage (in 

aqueous methylcellulose) for 60 days at doses as low as 8 mg/kg/day; renal and neurological effects were 

also reported (Knauf and Hobson 1979). All male SPF pigs fed 50 mg/kg/day (but not lower doses up to 

5 mg/kg/day) for 90 days exhibited depressed growth (beginning at week 4) prior to death between 

weeks 8 and 12 (Den Tonkelaar et al. 1978). 

In fetal and immature pups, hexachlorobenzene-induced changes body weight were observed only in the 

presence of maternal toxicity. No decreases in maternal or pup body weight were observed at doses as 

high as 100 mg/kg/day on gestation days 7–16 in CD-1 mice or as long as in utero and lifetime exposure 

of Sprague-Dawley rats to 2 mg/kg/day (Arnold et al. 1985; Courtney et al. 1976; Goldey and Taylor 

1992; Lilienthal et al. 1996; Taylor and Goldey 1990; Vos et al. 1979a, 1983). In Wistar rats, maternal 

and fetal body weight were decreased by gavage doses of at least 80 mg/kg/day of hexachlorobenzene in 

corn oil on gestation days 6–21 (Khera et al. 1974a); however, slight changes in the dosing protocols 

(fewer treatment days, use of aqueous gum tragacanth instead of corn oil) prevented weight loss. 

Likewise, Sprague-Dawley rat pup weight and viability were significantly reduced at birth, 5 days, and 

24 days, following lifetime dietary exposure of both male and female parent rats to at least 4 mg/kg/day 
HEXACHLOROBENZENE 95 

3. HEALTH EFFECTS 

(Grant et al. 1977). Maternal signs reported by these two studies (Grant et al. 1977; Khera et al. 1974a) 

included mortality, convulsions, and reduced fertility. 

3.2.2.3 Immunological and Lymphoreticular Effects 

Although epidemiological studies have suggested that consumption by mothers of fish high in 

organochlorines may affect the immune system of their infants, oral exposure of people to hexachlorobenzene has not been clearly associated with immunological effects. The levels of hexachlorobenzene, 

dieldrin, and p,p’-dichlorodiphenyldichloroethylene (p,p’-DDE) in milk samples collected shortly after 

the birth of Canadian Inuit infants correlated with increased risk of otitis media in their first year of life, 

but not with other serum immune parameters (immunoglobulins, cytokines, lymphocyte activation 

markers) (Dewailly et al. 2000). A similar study of immune parameters in umbilical blood of Canadian 

mothers with high or low consumption of fish from the Lower-North-Shore of the St. Lawrence River 

detected a statistically significant association only between decreased lymphocyte secretion of cytokine 

IL-10 and increased levels of hexachlorobenzene, p,p’-DDE, and mercury (Belles-Isles et al. 2000). 

However, evidence suggests that hexachlorobenzene may indirectly affect the immune system by 

inducing porphyria (see Section 3.2.2.2 Hepatic Effects). Mode-of-action studies of individuals with 

inherited and acquired porphyria (unrelated to hexachlorobenzene) have found that irradiated porphyrins 

may activate the complement system and stimulate mast cells and neutrophils to damage nearby tissues 

(Lim and Cohen 1999; Meola and Lim 1993). Additionally, an increased risk of porhyria cutanea tarda 

has been associated with HIV infection (reviewed Drobacheff et al. 1998; Egger et al. 2002). 

The effects of oral exposure to hexachlorobenzene to the immune system of animals appear to be speciesand strain- (Michielsen et al. 1997) dependent, with immunosuppression observed in mice (see below), 

monkeys (Iatropoulos et al. 1976) and bears (Bernhoft et al. 2000), and at least a partial stimulation of the 

immune system in rats (see below) and dogs (Gralla et al. 1977a). Additionally, a number of animal 

studies have observed inflammation and immune cell infiltration in tissues such as the liver (Arnold et al. 

1985; Ertürk et al. 1986), respiratory tract (Goldstein et al. 1978; Kimbrough and Linder 1974; Kitchin et 

al. 1982; Michielsen et al. 1997, 1999, 2001; Vos et al. 1979a, 1983), and skin (Koss et al. 1978; 

Michielsen et al. 1997, 2000; Schielen et al. 1993, 1995b; Torinkui et al. 1981) following oral exposure to 

hexachlorobenzene. The lowest dose to cause an immune response was 0.016 mg/kg/day, which induced 

peribiliary lymphocytosis in F1 Sprague-Dawley rats exposed for life to dietary doses (Arnold et al. 

1985). Because the mode-of-action is unclear, it is not known if these immune effects are secondary 

following toxicity to target organs or if they are involved in the etiology of disease in these organs. 
HEXACHLOROBENZENE 96 

3. HEALTH EFFECTS 

Mode-of-action studies in rats have suggested that the immune effects of hexachlorobenzene may be 

secondary to the accumulation of porphyrins produced by the liver in the spleen or other organs of 

immunological importance (Kennedy and Wigfield 1990; Kuiper-Goodman et al. 1977) or by the 

metabolic products of hexachlorobenzene (Schielen et al. 1995a). 

Immunosuppression has been observed in male mice following oral exposure to hexachlorobenzene. The 

most sensitive effects were seen at dietary doses as low as 0.6 mg/kg/day; 6 weeks of exposure 

significantly reduced spleen cell cytocidal activity (Loose et al. 1981) and 18 weeks of exposure 

decreased the mean survival time of mice following challenge with injected ascites tumor cells. Immune 

effects were seen as early as 3 weeks; feeding of 13 mg/kg/day decreased survival time following ascites 

tumor cell challenge (Loose et al. 1981), and feeding with 22 mg/kg/day decreased resistance to bacterial 

endotoxin (Salmonella typhosa lipopolysaccharide [LPS]) and to malarial infection (Plasmodium berghei) 

(Loose et al. 1978). Feeding at the same dose (22 mg/kg/day) for 6 weeks reduced primary and 

secondary plaque-forming cell responses to sheep red blood cells, reduced serum IgG, IgM, and IgA 

levels, increased spleen weight (Loose et al. 1977), increased susceptibility to infection by hepatis (but 

not cytomegalovirus or pneumonia) virus (Carthew et al. 1990), and reduced ability of spleen cells to lyse 

P388 tumor cells (Silkworth and Loose 1981). Feeding of 22 mg/kg/day for 37 (but not 13 or fewer) 

weeks of treatment decreased graft-versus-host response (measured by injecting harvested spleen cells 

into neonatal BDF1 mice) (Silkworth and Loose 1981). In addition to the effects seen in adult male mice, 

similar effects were observed in a developmental study. Pups of pregnant BALB/c mice fed doses as low 

as 0.5 mg/kg/day on gestation days 1–18 exhibited a marked, significant decrease in delayed type 

hypersensitivity (DTH) response when tested on postnatal day 45 (Barnett et al. 1987). Analyses of 

collected spleen cells found that 5 (but not 0.5) mg/kg/day decreased mixed lymphocyte response and 

decreased B cell numbers; neither dose affected spleen blastogenesis induced by T- or B-cell mitogens. 

In contrast to the immunosuppression observed in mice exposed to hexachlorobenzene, studies in rats 

have generally found stimulation of the immune system, as indicated by such effects as increased spleen 

and lymph node weights, increased neutrophil counts, and increased serum immunoglobulins. In feeding 

studies with male rats, exposure to at least 25 mg/kg/day of hexachlorobenzene for 3 weeks caused 

increased neutrophil counts, elevated popliteal lymph node weight with a corresponding proliferation of 

popliteal lymph node high-endothelial venules (Vos et al. 1979b), and spleen histopathology consisting of 

extramedullary hematopoiesis, enlarged marginal zones and follicles, and increased macrophage density 

in the marginal zones (Schielen et al. 1993; Vos et al. 1979b). In Wistar rats fed doses as low as 

50 mg/kg/day for 3 weeks, immunological effects observed included increased spleen and lymph node 
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3. HEALTH EFFECTS 

(popliteal, inguinal, and mediastinal, but not parathymic) weight; increased basophil, monocyte, and total 

leukocyte counts (Kennedy and Wigfield 1990; Koss et al. 1978; Schielen et al. 1995b; Vos et al. 1979a, 

1979b); increased splenic (but not lymph nodal) cell size; selectively activated B cell (but not T cell) 

subpopulations (Schielen et al. 1995b); increased splenic and serum levels of total IgM; and increased 

serum levels of specific IgM (anti-phosphatidylcholine IgM, and anti-single stranded DNA IgM) without 

increases in IgM against foreign antigens (tetanus toxoid, sheep erythrocytes, and bovine serum albumin) 

(Schielen et al. 1993, 1995b). These reported changes in B cell populations and IgM levels strongly 

suggest an autoimmune response. Remarkably, no changes were seen immune function tests for thymusindependent (Escherichia coli lipopolysaccharides) and thymus-dependent (tetanus toxoid) antibody 

responses, cell-mediated immunity (rejection of skin transplants, resistance to Listeria monocytogenes 

infection, or delayed-type hypersensitivity to Mycobacterium tuberculin), phagocytic competence 

(clearance of carbon particles), the mitogenic response of peripheral blood lymphocytes ( to pokeweed 

mitogen, concanavalin A, and phytohemagglutinin), or susceptibility to E. coli endotoxin in Wistar rats at 

doses up to 100 mg/kg/day for 3 weeks (Vos et al. 1979b). Feeding of male Wistar rats with doses as low 

as 8 mg/kg/day for 6 weeks reduced natural killer cell activity (Van Loveren et al. 1990). Atrophy of the 

thymus was observed in male rats fed 100 mg/kg/day for 3 weeks (Vos et al. 1979b) and in female Wistar 

rats fed 15 mg/kg/day for 13 weeks (Schielen et al. 1995a). 

Two developmental studies in rats by the same laboratory reported hexachlorobenzene effects on humoral 

and cellular immunity (Vos et al. 1979a, 1983). The first study reported that hexachlorobenzene exposure 

strongly enhanced humoral immunity (antibody response to tetanus toxoid), but slightly depressed 

cellular immunity (as evaluated by susceptibility to infection, skin graft rejection time, and response to 

mitogens) in pups whose mothers were fed 2.5 or 7.5 mg/kg/day of hexachlorobenzene from early 

pregnancy and continuing through gestation and lactation; after weaning pups were fed the same doses for 

an additional 2 weeks prior to testing (Vos et al. 1979a). In both treatment groups, resistance to infection 

(L. monocytogenes and Trichinella spiralis) was reduced, the IgG response to tetanus toxoid was 

significantly increased, and histopathology found proliferation of high-endothelial venules in the 

paracortex of the lymph nodes. At 7.5 mg/kg/day, the IgG response to Trichinella was increased, 

increases were seen in blood levels of eosinophils, basophils, IgM and IgG, and histopathology found 

accumulation of foamy macrophages in the lung and increased extramedullary hematopoiesis in the 

spleen. Neither dose affected rejection of skin transplant, passive cutaneous anaphylaxis, IgM response to 

LPS, mitogenic responsiveness of lymphocytes, clearance of carbon particles, or clearance of Listeria 

monocytogenes. However, the follow-up study found that hexachlorobenzene stimulated both humoral 

and cell-mediated immunity in the Wistar rat pups whose mothers were fed 0.2, 1, or 5 mg/kg/day of 
HEXACHLOROBENZENE 98 

3. HEALTH EFFECTS 

hexachlorobenzene from early pregnancy through lactation; weaning pups were fed the same doses up to 

7 months (Vos et al. 1983). Cytotoxicity of spleen cells (to injected lymphoma cells) was not affected by 

treatment. Treatment with at least 0.2 mg/kg/day significantly increased the IgG and IgM response to 

tetanus toxoid and delayed-type hypersensitivity reaction to ovalbumin, and induced accumulation of 

foamy macrophages in the lungs. At 1 mg/kg/day and above, increased popliteal lymph node weight with 

a corresponding increased in lymph node cellular proliferation were observed, and cell proliferation was 

also detected in the endothelial cells lining pulmonary capillaries and venules. At 5 mg/kg/day, increases 

were observed in spleen, lung, and mesenteric lymph node weight; serum IgM (but not IgG) levels; the 

relative number of basophils in the blood; and extramedullary hematopoiesis of the spleen. These effects 

demonstrated stimulation of both humoral and cell-mediated immunity. 

A strain-dependent correlation between immunological and dermal effects was observed in female rats 

(Michielsen et al. 1997). For 28 days, Wistar rats were fed 22 or 45 mg/kg/day, while Lewis rats were 

fed 8 or 22 mg/kg/day, and Brown Norway rats were fed 8, 22, or 45 mg/kg/day of hexachlorobenzene. 

Brown Norway rats were the most sensitive to hexachlorobenzene exposure, and correlations were 

observed between the incidence and severity of immune responses and dermal lesions. In contrast, Wistar 

rats were the most resistant, and correlations were not apparent, while some correlations were seen in 

Lewis rats. Hexachlorobenzene induced skin lesions in all treatment groups, most severe in Brown 

Norway rats and least severe in Wistar rats, characterized by epidermal hyperplasia, inflammatory 

infiltrate in the dermis, and activation (due to hypertrophy and proliferation) of endothelial cells in dermal 

vessels. Relative spleen weights were significantly increased in a dose-related fashion in all three strains 

while popliteal lymph node weight was increased in the high-dose Lewis and Brown Norway rats, but not 

in Wistar rats. All strains showed increases in IgM, but only Brown Norway rats exhibited increases in 

serum IgE and IgG. However, lung pathology was not strain dependent; histopathology observed lung 

lesions consisting of venules lined with unusually plump endothelial cells and surrounded by large 

perivascular infiltrate and accumulation of alveolar macrophages. The authors concluded that the 

inflammatory responses in the skin and lungs were of different etiologies, and speculated an involvement 

of the immune system in the observed dermal lesions. 

Studies of the immunological effects of hexachlorobenzene in nonrodents observed evidence of 

immunosuppression in monkeys and bears, and stimulation of the immune system in dogs. In female 

Rhesus monkeys, gavage with hexachlorobenzene in methylcellulose at doses as low as 8 mg/kg for 

60 days caused thymic cortical atrophy, consisting of a reduction or absence of individual lobules, 

increased numbers of thymic corpuscles, and medullar hyperplasia or reticular cells, macrophages, plasma 
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3. HEALTH EFFECTS 

cells, and lymphocytes. However, the use of only one or two monkeys in this study diminishes the 

reliability of these data (Iatropoulos et al. 1976). A significant correlation between hexachlorobenzene 

levels and decreased IgG was observed in an analysis of sera from 56 polar bears in Svalbard, Norway, 

observed (Bernhoft et al. 2000). Because similar correlations were also observed for three 

polychlorinated biphenyl congeners (99, 194, and 206), this effect cannot be clearly attributed to hexachlorobenzene. Although no immunologically-related gross pathology or histopathology were observed 

in a 21-day oral study in female dogs at doses as high as 150 mg/kg/day (Sundlof et al. 1981), nodular 

hyperplasia of gastric lymphoid tissue was found in all beagle dogs given hexachlorobenzene in gelatin 

capsules daily for 12 months at doses as low as 10 mg/kg/day and an increased incidence of infebrile 

neutrophilia (increased numbers of blood neutrophils without fever) was observed at 10 and 

110 mg/kg/day (Gralla et al. 1977a). 

The highest NOAEL values and all reliable LOAEL values for immunologic effects in each species and 

duration category are recorded in Table 3-2 and plotted in Figure 3-2. 

3.2.2.4 Neurological Effects 

The evidence of neurotoxicity in humans following oral exposure to hexachlorobenzene was provided by 

studies of people in southeast Turkey who consumed contaminated bread in the late 1950s. The ingested 

dose of hexachlorobenzene by exposed persons was estimated to be in the range of 0.05–0.2 g/day 

(0.7–2.9 mg/kg/day for a 70-kg person) during the episode (Cam and Nigogosyan 1963). 

Neurological symptoms included loss of appetite, tremors, convulsions, and “weakness that often made it 

impossible to eat with a knife and fork, rise from a squat, or climb stairs” (Gocmen et al. 1989; Peters et 

al. 1982). Follow-up studies of 25 (Peters et al. 1982) and 30 years (Cripps et al. 1984) included 161 and 

204 patients, respectively. They found that neurological symptoms persisted in adults who had been 

exposed as children, and included weakness (62–66%), paresthesia (spontaneous tingling or burning 

sensations, 55%), sensory shading (graded sensory loss that diminishes upon testing more proximally and 

is indicative of polyneuropathy, 61–63%), myotonia (delayed muscle relaxation after an initial 

contraction, 38–50%), and cogwheeling (irregular jerkiness of movement due to increased muscle tone as 

seen in Parkinson’s disease, 29–41%). During the grain poisoning epidemic, there was an extremely high 

(95%) rate of mortality in infants under 2 years of age, who had been breast fed by mothers who had 

ingested the contaminated bread; these children exhibited convulsions, tremors, and progressive weakness 

prior to death (Cripps 1990; Peters et al. 1966). Analysis of human milk from exposed women and 
HEXACHLOROBENZENE 100 

3. HEALTH EFFECTS 

unexposed controls in this epidemic showed hexachlorobenzene concentrations of 0.51 and 0.07 ppm, 

respectively (Cripps et al. 1984; Gocmen et al. 1989; Peters et al. 1982, 1987). A study investigating the 

potential effects of consuming fish from the Great Lakes was unable to correlate hexachlorobenzene 

levels in umbilical blood or breast milk with infant intelligence test results (Darvill et al. 2000). 

Multiple studies have shown that hexachlorobenzene induces serious neurological effects such as 

convulsions, tremors (intermittent and constant), lethargy, and progressive weakness in rats, mice, rabbits, 

pigs, monkeys, and quail (Cabral et al. 1979; Cripps 1990; De Matteis et al. 1961; Den Besten et al. 1993; 

Den Tonkelaar et al. 1978; Grant et al. 1977; Hahn et al. 1988; Kennedy and Wigfield 1990; Kimbrough 

and Linder 1974; Khera 1974a; Knauf and Hobson 1979; Ockner and Schmid 1961; others). In several 

studies, these effects were only seen prior to death or in treatment groups with significant mortality. The 

lowest dose to cause such serious effects was 8 mg/kg/day in Sprague-Dawley rats as part of a 

multigeneration study, with convulsions preceding death (Grant et al. 1977). In comparison, no tremors 

were seen in female Agus or Wistar rats fed diets containing 5 mg/kg/day of hexachlorobenzene for 

75–90 weeks (Smith and Cabral 1980). Similar effects have been seen in other species. Mice orally 

exposed to 26 mg/kg/day of hexachlorobenzene for up to 17 weeks exhibited severe tremors prior to 

death (Hahn et al. 1988). Male SPF pigs fed 50 mg/kg/day of hexachlorobenzene for 90 days exhibited 

tremors, panting, and unsteady gait without histopathology (Den Tonkelaar et al. 1978). Adult female 

Rhesus monkeys given oral doses of hexachlorobenzene for 60 days suffered severe tremors and muscular 

weakness at doses as low as 64 mg/kg/day and marked lethargy and weakness were observed at 

128 mg/kg/day (Knauf and Hobson 1979). Two of three infant Rhesus monkeys whose mothers were 

treated with 64 mg/kg/day for up to 60 days during lactation displayed hypoactivity, lethargy, and ataxia, 

and subsequently died (Iatropoulos et al. 1978). The offspring of pregnant rats feed 200 (but not 

5) mg/kg/day for an unspecified period during gestation exhibited convulsions, tremors, and progressive 

weakness (Cripps 1990). 

Other studies have investigated more subtle neurological effects resulting from oral exposure to hexachlorobenzene. Electrophysiological changes (dysrhythmic electroencephalogram) in the central nervous 

system were demonstrated in dogs receiving doses of $50 mg/kg/day for 21 days (Sundlof et al. 1981). 

Another functional experiment observed axonal effects in the sciatic nerve (fibrillations, repetitive or 

pseudomyotonic discharges, and mild slowing of conduction velocities) in rats fed 40 mg/kg/day of hexachlorobenzene for 20 weeks or at least 3.75 mg/kg/day for 2 years (Sufit et al. 1986). In male Wistar rats 

gavaged with doses as low as 317 mg/kg/day for 4 weeks, significant decreases were observed in the rate 

of T4 uptake into cerebrospinal fluid and brain tissue (van Raaij et al. 1994). Additionally, gavage 
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3. HEALTH EFFECTS 

treatment of Chbb THOM and Wistar rats with 1,000 mg/kg/day of hexachlorobenzene in a water-Tween 

suspension for up to 28 days induced changes in brain phospholipids (phosphatidylcholine, 

phosphatidylethanolamine, phosphatidylinositiol, phosphatidylserine, and sphingomyelin) without 

affecting brain porphyrin levels (Billi de Catabbi et al. 2000b; Cochon et al. 2001). Because these effects 

were seen prior to the onset of porphyrin accumulation in the liver, the authors concluded that the effects 

of hexachlorobenzene on phospholipids in the brain were different from its effects on phospholipids of 

the liver or Harderian gland, which reportedly occur following porphyrin accumulation. 

One of two developmental studies in rats that investigated the neurobehavioral effects of hexachlorobenzene observed hyperactivity. Two weeks prior to mating, female Sprague-Dawley rats were gavaged 

for 4 days with 2.5 or 25 mg/kg/day (Goldey and Taylor 1992; Taylor and Goldey 1990). In the first 

3 weeks postnatally, both treatment groups of pups exhibited a significantly increased level of hyperactivity compared to controls. Specifically, treated pups reoriented themselves significantly more quickly 

in a negative geotaxis test, required less time in an olfactory discrimination test, and demonstrated 

increased exploratory activity in a motor activity test (postnatal days 15–20). No significant effects on 

learning (swim T-maze) or motor activity (measured in older offspring on postnatal days 40 and 50, 

respectively) were detected. Pups exposed to 25 mg/kg/day exhibited decreased acoustic startle response 

(ASR) on postnatal day 23 and increased ASR on postnatal day 90. Another study fed female Wistar rats 

hexachlorobenzene for 90 days prior to mating through lactation and offspring were maintained on the 

same diet until postnatal day 150 (Lilienthal et al. 1996). In pups exposed to 1.3 mg/kg/day (but not 

lower doses up to 0.64 mg/kg/day), significant decreases were seen in operant learning 

(“post-reinforcement pause” and “index of curvature”) on postnatal day 150. However, because the rats 

were exposed both developmentally and as adults, the development significance of changes in operant 

learning is unclear. No changes were seen in an open field activity test (a measure of early locomotor 

skills) on day 21 or an active avoidance learning test on day 90. 

The highest NOAEL values and all reliable LOAEL values for neurological effects in each species for 

intermediate duration category are shown in Table 3-2 and plotted in Figure 3-2. 
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3. HEALTH EFFECTS 

3.2.2.5 Reproductive Effects 

Epidemiological studies suggest that hexachlorobenzene may cause spontaneous abortion in women. In 

Southeastern Turkey, consumption of bread made from grain treated with hexachlorobenzene resulted in 

widespread poisoning between 1955 and 1959. Although no quantitation of exposure was presented in 

any of these clinical reports, an estimated dose of 0.05–0.2 g/day (0.7–2.9 mg/kg/day for a 70-kg person) 

is considered to be reliable by the original investigators of the Turkey epidemic (Cam and Nigogosyan 

1963). A follow-up study conducted between 1977 and 1981 identified 42 porphyric mothers who had 

been exposed as children, with 188 pregnancies (Peters et al. 1982, 1987). Of these, 15 were fetal deaths 

(13 miscarriages and 2 stillbirths), and 31 produced children who died in the first several years of life. 

Similarly, another follow-up study conducted 20–30 years after initial exposure identified 57 porphyric 

mothers, who had a total of 276 pregnancies (Gocmen et al. 1989). Of these, 23 were fetal deaths, and 

54 produced children who died in the first several years of life. Porphyric mothers had an average of 

0.51 ppm hexachlorobenzene in their breast milk, compared to 0.07 ppm in unexposed controls (Gocmen 

et al. 1989). The degree to which these two exposed populations overlap and the expected frequencies of 

adverse pregnancy outcomes for the unexposed cohorts were unclear. Surviving offspring of porphyric 

mothers were clinically normal, and had urine and stool porphyrin levels similar to control children. 

A subsequent retrospective study, conducted 40 years after initial exposure, compared three groups of 

42 women (controls from outside the exposed area and women from the hexachlorobenzene-exposed 

region either with or without a diagnosis of porphyria cutanea tarda) (Jarrell et al. 1998). The incidence 

of women with blood levels of hexachlorobenzene exceeding 1 ng/mL was greater in women with 

porphyria cutanea tarda or women living in the contaminated region than country-wide controls and 

correlated (across exposure-groups) with an increased risk of spontaneous abortion. Notably, blood levels 

did not correlate with the number of pregnancies, sex ratio of born children, or onset of menopause. 

Statistically significant increases in the levels of inhibin (a hormone secreted by ovarian granulosa cells to 

decrease the release of follicle-stimulating hormone [FSH] from the pituitary) were observed in women 

diagnosed with porphyria cutanea tarda. Because no exposure-related differences were seen for FSH or 

estradiol, the biological significance is unclear, but ovarian effects would be consistent with animal 

studies (see below). 

Studies of other populations with exposure to multiple organochlorines did not find significant 

differences in blood hexachlorobenzene levels between controls and cases of spontaneous abortion in 

Italy (Leoni et al. 1986, 1989) or Germany (Gerhard et al. 1998). Average maternal blood levels of 
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3. HEALTH EFFECTS 

hexachlorobenzene were 1.6 and 0.679 ng/mL, respectively. Similarly, no changes in reproductive 

outcomes in Xixin, China, were detected following the cessation of agricultural uses of hexachlorobenzene (Huang et al. 1989). 

For hexachlorobenzene, animal studies have identified adverse effects in the ovaries at doses as low as 

0.01 mg/kg/day (Alvarez et al. 2000; Babineau et al. 1991; Bourque et al. 1995; Foster et al. 1992a, 

1992b, 1995a, 1995b; Jarrell et al. 1993; Lecavalier et al. 1994; Iatropoulos et al. 1976; Knauf and 

Hobson 1979; MacPhee et al. 1993; Muller et al. 1978; Sims et al. 1991), in the testes at doses as low as 

10 mg/kg/day (Gralla et al. 1977a; Smith et al. 1985), and on reproductive function with doses of at least 

16 mg/kg/day (Grant et al. 1977; Simon et al. 1979). 

The ovaries are a sensitive target organ for hexachlorobenzene. Distribution studies have identified the 

ovaries as a site of hexachlorobenzene accumulation (Foster et al. 1993; Sitarska et al. 1995; others). 

Studies have reported changes in organ weight; histological (light microscopy) and ultrastructural 

(electron microscopy) degenerative changes; and altered serum levels of gonadal hormones (estrogen and 

progesterone). Investigations into the mode-of-action generally found disruptions in steroidogenesis. 

The acute data for ovarian effects are limited (Foster et al. 1993; Lecavalier et al. 1994). Mild 

morphological changes in the ovary were detected in female Sprague-Dawley rats given a single gavage 

dose of at least 400 mg/kg in corn oil (Lecavalier et al. 1994). Serum progesterone levels were increased 

in female Sprague-Dawley rats that were superovulated prior to gavage treatment with 50 mg/kg/day of 

hexachlorobenzene in corn oil for 5 days but not in normally-cycling rats similarly treated (Foster et al. 

1993). 

A 90-day assay in adult female Cynomolgus monkeys observed ovarian toxicity at the lowest dose tested, 

0.001 mg/kg/day of hexachlorobenzene (Bourque et al. 1995). Ultrastructural analyses of developing ova 

detected mitochondrial changes, which increased in frequency and severity with dose. Swelling of the 

cristae resulting in abnormal intracristae spaces was seen at 0.01 mg/kg/day; mitochondrial matrices 

became more coarsely granular and exhibited occasional irregular morphology at 0.1 mg/kg/day; and 

mitochondria had “electron-lucent” matrices and reduced membrane integrity at 10 mg/kg/day. A similar 

increase in frequency and severity was observed for lesions in follicular cells: abnormal nuclei were seen 

in “a few cells” at 0.01 mg/kg/day, while nuclear membrane infolding was clearly apparent at 

0.1 mg/kg/day. Abnormal spaces between follicular cells were observed at 1 mg/kg/day, and follicular 

cells exhibited abnormal lipid accumulation and deeply folded and indented nuclear membranes at 
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3. HEALTH EFFECTS 

10 mg/kg/day. Thecal cells exhibited deformed nuclei only at 10 mg/kg/day. Since mitochondrial 

changes may represent nonspecific cell injury, the specific mode of action of degenerative ovarian 

changes remains unclear. The LOAEL of 0.01 mg/kg/day from this study was used to calculate an 

intermediate oral MRL of 0.0001 mg/kg/day as described in the footnote to Table 3-2 and in Appendix A. 

Supporting data are provided by previous 90-day studies in female Cynomolgus monkeys that had doserelated degenerative changes in oocytes and ovaries at all doses tested, 0.1–10 mg/kg/day (Babineau et al. 

1991; Foster et al. 1992a, 1995a; Jarrell et al. 1993; Sims et al. 1991). Increasing in frequency and 

severity, oocyte effects progressed from increased numbers of lysosomal elements and irregularly 

arranged thecal layer cells to altered oocyte morphology (shape irregularities, increased granularity and 

density, less distinct membrane), cytoplasmic vacuolation, lysomal aggregation, and pyknosis of follicular 

granulosa cells. Similarly, ovarian changes progressed from cellular hypertrophy, increasing 

columnarization of normally cuboidal cells, and small lipid inclusions, to cellular necrosis and separation 

of epithelium from connective tissue, stratification and elongation of epithelial cells, and increased 

numbers of lysosomes, vesicles, and lipid inclusions. Additionally, luteal phase progesterone levels were 

reduced at 1 mg/kg/day, ovulatory surge estrogen levels were decreased at 10 mg/kg/day, and menstrual 

cycle length variability was increased at 10 mg/kg/day (Foster et al. 1992a, 1995a). Remarkably, doses 

up to 10 mg/kg/day did not affect fertility (measured in oocytes by in vitro tests), serum inhibin levels, or 

numbers or size of oocytes, follicles, or corpora lutea. Notably, liver lesions (cholestasis, accentuated 

zonation, increased portal density, and mid-zone vacuolation) were only seen at the higher doses of 1 and 

10 mg/kg/day. 

Earlier studies in female Rhesus monkeys observed similar lesions. In female Rhesus monkeys given 

hexachlorobenzene for 60 days, ovarian effects seen at 8 mg/kg/day included cortical degeneration, 

reduced numbers of primary follicles with a concurrent increase in relative corpora lutea volume, multiple 

follicular cysts, and a thickening of the ovarian germinal epithelium with cells exhibiting a columnar 

appearance progressing to pseudostratification (Iatropoulos et al. 1976). These effects increased in 

incidence and severity with dose. At 32 mg/kg/day and above, epithelial nuclei were pyknotic 

(condensed) and karyorrhectic (fragmented) and at 128 mg/kg/day, ovarian cortices were predominated 

by dense stroma. These ovarian changes were similar to those normally seen in menopause, and indicate 

that the corpora lutea were not producing steroids. A subsequent study in female Rhesus monkeys found 

that serum cholesterol was increased by gavage doses of 8 mg/kg/day of hexachlorobenzene in methyl 

cellulose for 60 days (Knauf and Hobson 1979); this effect may be secondary to changes in ovarian 

steroidogenic activity. These findings are supported by a study in which hexachlorobenzene blocked 
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ovulation (estrogen and progesterone remained low, LH and FSH continued to climb, and menstruation 

was delayed) in one of four female Rhesus monkeys gavaged with 4 mg/kg/day of hexachlorobenzene in 

aqueous methyl cellulose for up to 78 days (Muller et al. 1978). A lower-dose study found no changes in 

serum levels of estrogen, progesterone, FSH, or LH in female Rhesus monkeys fed 0.03 mg/kg/day of 

hexachlorobenzene in monkey chow for 11 months (Rozman et al. 1978). The difference between this 

study and the one in Cynomolgus monkeys that observed changes in estrogen and progesterone levels 

(Foster et al. 1992a, 1995a) may reflect strain specificity or differences in absorption following disparate 

methods of oral exposure. 

No evidence of histopathology was found in the ovaries of female beagle dogs given hexachlorobenzene 

in gelatin capsules with corn oil at doses up to 100 mg/kg/day for 21 days; data for other reproductive 

organs were not reported (Sundlof et al. 1981). 

Although rat studies have not investigated doses as low as those used in monkey studies, they have 

identified histological evidence of degeneration and ultrastructural changes in the ovaries of animals 

treated with hexachlorobenzene. Gavage treatment of female Wistar rats with 1,000 mg/kg/day of 

hexachlorobenzene for 30 days caused degenerative lesions (increased numbers of atresic follicles, 

inflammatory infiltration of primary follicles, stratification and proliferation of ovarian surface epithelial 

cells, and irregular nuclei in epithelial cells) and changes in hormone and hormone receptor levels 

(decreased serum estradiol and prolactin, increased FSH, decreased estrogen receptor levels) (Alvarez et 

al. 2000). Similarly, a 21-day study detected increased serum progesterone levels in female SpragueDawley rats gavaged with at least 1 mg/kg/day of hexachlorobenzene (Foster et al. 1992b). In treated 

animals, the number of ova produced per rat decreased significantly and the length of estrus increased 

significantly. Another study in female Sprague-Dawley rats observed suggestive evidence of ovarian 

lesions (increased prominence of Golgi complexes, smooth endoplasmic reticulum, and free polysomes) 

at 10 (but not 1 or 100) mg/kg/day (MacPhee et al. 1993). The differences observed between rats and 

monkeys for changes in progesterone levels may be related to differences in their cycle lengths. 

To investigate the contributions of adrenal steroidogenesis, adult female Sprague-Dawley rats were 

ovariectomized prior to treatment (Foster et al. 1995a). Gavage treatment for 30 days at doses as low as 

1 mg/kg/day decreased serum corticosterone, while serum cortisol was decreased only at 100 mg/kg/day; 

no effects were seen on aldosterone or progesterone levels. The authors concluded that hexachlorobenzene induced alterations in steroidogenesis of cells of the inner zone of the adrenal cortex. 
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Some intermediate-duration experiments have demonstrated that hexachlorobenzene adversely affects 

reproductive performance. In a multigenerational study in which male and female Sprague-Dawley rats 

were fed hexachlorobenzene at doses of 0.5–32 mg/kg/day through premating and two series of mating, 

gestation, and lactation for up to four generations, statistically significant decreases in fertility and 

increases in the number of stillborns were observed at doses as low as 16 mg/kg/day and average litter 

size was decreased at 8 mg/kg/day and above (Grant et al. 1977). A study which dosed only male rats 

were dosed with 70 mg/kg/day of hexachlorobenzene for 5 consecutive days prior to 14 weeks of 

consecutive mating found decreased male reproductive performance (mating index), but no changes in 

fertility (numbers of inseminated females made pregnant) (Simon et al. 1979). In contrast, in two studies 

in which only females were dosed, no reproductive toxicity was observed, either in female SpragueDawley rats fed 7 mg/kg/day continuously from 96 days prior to first mating through gestation of two 

successive litters (Kitchin et al. 1982) or in female Cynomolgus monkeys given 10 mg/kg/day orally for 

90 days (Jarrell et al. 1993). 

Similarly, no reproductive toxicity was observed in a chronic assay in which both male and female 

Sprague-Dawley rats were fed doses of hexachlorobenzene up to 2 mg/kg/day from 3 months prior to 

mating through weaning (Arnold et al. 1985). 

The other chronic data available are pertinent to testicular damage (Gralla et al. 1977a; Smith et al. 1985). 

In male Fischer 344/N rats fed 10 mg/kg/day of hexachlorobenzene in arachis oil for 90 weeks, testicular 

weight was significantly increased and testicular interstitial cell tumors were more severe (although 

incidence was not affected) compared to controls (Smith et al. 1985). Slight testicular degeneration, with 

numerous spermatogonic giant cells and incomplete complement of spermatogonia in the seminiferous 

tubules, was observed in two of six male beagle dogs given 110 mg/kg/day of hexachlorobenzene in 

gelatin capsules with corn oil for 12 months (Gralla et al. 1977a). Additionally, retarded sexual 

maturation of the testes was observed in male SPF pigs fed 50 (but not lower doses up to 5 mg/kg/day of 

hexachlorobenzene for 90 days (Den Tonkelaar et al. 1978). 

The highest NOAEL values and all reliable LOAEL values for reproductive effects in each species and 

duration category are shown in Table 3-2 and plotted in Figure 3-2. 
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3. HEALTH EFFECTS 

3.2.2.6 Developmental Effects 

Human and animal studies have demonstrated that hexachlorobenzene crosses the placenta to accumulate 

in fetal tissues; additionally, hexachlorobenzene is concentrated in milk and can be transferred to the 

suckling neonate (for more information, see Sections 3.4.1 and 3.4.2). 

Oral exposure to hexachlorobenzene has been associated with serious developmental toxicity in a study of 

a poisoning epidemic (Cripps et al. 1984; Gocmen et al. 1989; Peters et al. 1966, 1982, 1987). In 

southeast Anatolia, Turkey, ingestion of an estimated 0.7–2.9 mg/kg/day of hexachlorobenzene between 

1955 and 1959 (in bread made from grain treated with hexachlorobenzene as a fungicide) resulted in 

dramatic developmental effects, including a 95% mortality rate in infants under 2 years of age who had 

been breast fed by exposed mothers (Peters et al. 1966). These infants were diagnosed with a condition 

known as pembe yara or "pink sore" because of associated skin lesions consisting of blistering, 

epidermolysis, and annular erythema. The cause of death in these infants was cardiorespiratory failure; 

weakness and convulsions were also seen frequently. Older children, between the ages of 6 and 15 years, 

exhibited a condition known as kara yara or “black sore” more frequently than younger children or 

adults. The symptoms of this disease began with photosensitivity and progressed within 6 months to 

include hyperpigmentation, dermal fragility (resulting in ulcerating lesions and severe mutilating scars), 

and hirsutism (Cam and Nigogosyan 1963; Dogramaci 1964; Gocmen et al. 1989). Mortality was 10% 

among kara yara patients. These skin lesions (pink sore and black sore) have been diagnosed as 

porphyria cutanea tarda, a specific type of vesiculobullous porphyria. The porphyrias are a class of 

inherited and acquired diseases caused by enzymatic defects in heme biosynthesis, leading to the 

generation of porphyrins, which may cause tissue damage, especially in the skin (for more information, 

see Sections 3.2.2.2, Hepatic Effects and Dermal Effects, and 3.5.2, Mechanisms of Toxicity). Similar 

dermal lesions, but no increase in mortality incidence, were reported for exposed adults, who also 

exhibited neurological disorders (weakness and diminished muscle control). 

Follow-up studies have found persistent symptoms of developmental toxicity in a cohort of 252 adults 

(162 men and 90 women) who had been exposed as children in the poisoning epidemic (average age of 

the cohort at the time of exposure was 7.6 years) (Cripps et al. 1984; Gocmen et al. 1989; Peters et al. 

1982). Short stature was seen in 42.1% of the patients, considered striking in comparison to their 

unexposed siblings. Additionally, 66.6% of the exposed patients exhibited osteoporosis of bones in the 

hands, associated with distinctive small hands, painless swelling, and spindling of fingers. Osteoporosis 

and osteosclerosis have been observed in adult hexachlorobenzene-exposed rats (see Section 3.2.2.2, 
HEXACHLOROBENZENE 108 

3. HEALTH EFFECTS 

Musculoskeletal Effects). In addition to the profound weakness and decreased muscle control observed in 

exposed adults (see Section 3.2.2.4, Neurological Effects), this cohort also presented parestheisa 

(spontaneous tingling or burning sensations, in 53.6% of patients) and graded sensory loss indicative of 

polyneuropathy (in 60.6% of patients). 

Two follow-up investigations have found potential reproductive effects in women exposed as children to 

hexachlorobenzene in the Turkish epidemic (Gocmen et al. 1989; Jarrell et al. 1998). A study conducted 

20–30 years after initial exposure identified 57 porphyric mothers with 23 fetal deaths and 54 children 

who died in the first several years of life from a total of 276 pregnancies (Gocmen et al. 1989). Porphyric 

mothers had an average of 0.51 ppm hexachlorobenzene in their breast milk, compared to 0.07 ppm in 

unexposed controls. The results of this study were inconclusive because no information was provided 

regarding the expected incidence of fetal deaths and newborn-death in this population. Another study 

found a statistically significant increased risk of abortion among a subset of exposed women who 

exhibited porphyria cutanea tarda and had blood levels of hexachlorobenzene above 1.0 ng/mL (Jarrell et 

al. 1998). 

A German case-control study found that adipose hexachlorobenzene levels in 18 male patients who 

underwent orchidopexy to correct unilateral or bilateral undescended testis (mean age 4.2 years) were 

3-fold higher compared to a group of 30 male control patients (mean age 3.5 years); this difference was 

highly significant statistically (Hosie et al. 2000). These results were inconclusive because a similar 

correlation was also observed for heptachloroepoxide (but not for other organochlorines measured). This 

study is also limited by its small study size and lack of age-adjustment. 

Although epidemiological studies have suggested that consumption by mothers of fish high in 

organochlorines may affect the immune system of their infants, perinatal exposure to hexachlorobenzene 

has not been clearly associated with immunological effects. The levels of hexachlorobenzene, dieldrin, 

and p,p’-dichlorodiphenyldichloroethylene (p,p’-DDE) in milk samples collected shortly after the birth of 

Canadian Inuit infants correlated with increased risk of otitis media in their first year of life, but not with 

other serum immune parameters (immunoglobulins, cytokines, lymphocyte activation markers) (Dewailly 

et al. 2000). A similar study of immune parameters in umbilical blood of Canadian mothers with high or 

low consumption of fish from the Lower-North-Shore of the St. Lawrence River detected a statistically 

significant association only between decreased lymphocyte secretion of cytokine IL-10 and increased 

levels of hexachlorobenzene, p,p’-DDE, and mercury (Belles-Isles et al. 2000). 
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Acute-duration developmental studies have verified that hexachlorobenzene impaired neurological 

development at doses as low as 2.5 mg/kg/day in rats (Goldey and Taylor 1992) and produced teratogenic 

abnormalities at doses as low as 40 mg/kg/day (Courtney et al. 1976; Khera 1974a). Intermediateduration developmental studies in rats have seen immunodevelopmental effects at 0.5 mg/kg/day (Barnett 

et al. 1987); neurodevelopmental effects at 1.3 mg/kg/day (Lilienthal et al. 1996); and reduced neonatal 

viability and growth, and organ weight changes at 4–5 mg/kg/day (Grant et al. 1977; Kitchin et al. 1982; 

Vos et al. 1979a, 1983). In Rhesus monkey pups, death (accompanied by neurological effects, lung 

edema, and liver damage) resulted from nursing for 15–38 days from female monkeys exposed to 

64 mg/kg/day of hexachlorobenzene (Bailey et al. 1980; Iatropoulos et al. 1978). One chronic study was 

found (Arnold et al. 1985). In a multigenerational reproductive study in Sprague-Dawley rats, 

hexachlorobenzene induced minor liver effects (peribiliary lymphocytosis and fibrosis) at doses as low as 

0.016 mg/kg/day in F1 adults and induced more serious effects (decreased F1 neonatal viability; tumor 

formation in F1 adults) at doses as low as 2 mg/kg/day of hexachlorobenzene. 

The most sensitive acute study evaluated neurodevelopmental end points and detected evidence of 

hyperactivity in Sprague-Dawley rat pups (Goldey and Taylor 1992; Taylor and Goldey 1990). This 

study was considered acute because virgin female Sprague-Dawley rats were gavaged for 4 days with 

2.5 or 25 mg/kg/day of hexachlorobenzene 2 weeks prior to mating. Compared to controls, pups from 

both treatment groups reoriented themselves significantly more quickly in a negative geotaxis test, 

required less time in an olfactory discrimination test (postnatal days 6, 8, and 10), and demonstrated 

increased exploratory activity in a motor activity test (postnatal days 15–20). Pups exposed to 

25 mg/kg/day exhibited decreased acoustic startle response (ASR) on postnatal day 23 and increased ASR 

on postnatal day 90. The LOAEL of 2.5 mg/kg/day from this study has been used to calculate an acute 

oral MRL of 0.008 mg/kg/day as described in the footnote to Table 3-2 and in Appendix A. 

Other acute studies have used higher doses to investigate traditional end points of developmental toxicity. 

An acute single-dose study found an increase in the overall incidence of fetal abnormalities (but not any 

specific abnormality) in the fetuses of pregnant female CD-1 mice gavaged with 100 mg/kg/day of 

hexachlorobenzene on gestation days 7–16; cleft palate and renal agenesis were the most common 

anomalies noted (Courtney et al. 1976). Three acute (3–10 days) and one intermediate (15 days) 

developmental toxicity experiments in pregnant Wistar rats observed increases in the incidences of sternal 

variations and the 14th rib formation at $40 mg/kg/day (Khera 1974a). At $80 mg/kg/day, decreased fetal 

and maternal body weights were seen with other maternal effects. 
HEXACHLOROBENZENE 110 

3. HEALTH EFFECTS 

The most sensitive intermediate-duration study evaluated immunodevelopmental toxicity. Pups of 

pregnant BALB/c mice fed doses as low as 0.5 mg/kg/day on gestation days 1–18 exhibited a marked, 

significant decrease in delayed type hypersensitivity (DTH) response when tested on postnatal 

day 45 (Barnett et al. 1987). Splenic effects (decreased B cell numbers and decreased mixed lymphocyte 

responses) were seen at 5 mg/kg/day (but not 0.5 mg/kg/day). Studies in Wistar rats have also 

demonstrated immunodevelopmental toxicity. In pups born to dams exposed through gestation and 

lactation to hexachlorobenzene and continued on the same diets, 0.2 mg/kg/day, increased immune 

responses (IgM and IgM responses to tetanus toxin, delayed-type reaction to ovalbumin, and pulmonary 

accumulation of foamy macrophages) were noted (Vos et al. 1983); exposure to 2.5 mg/kg/day decreased 

resistance to infection and caused lymph node endothelial proliferation. Additional signs of 

developmental toxicity were seen in these pups at higher doses: pup mortality, liver weight, and popliteal 

lymph node weights were increased at 5 mg/kg/day; and adrenal weight and serum immunoglobulin 

concentrations (IgM and IgG) were increased at 7.5 mg/kg/day (Vos et al. 1979a, 1983). Similar 

developmental effects were observed in an intermediate-duration reproductive study which fed female 

Sprague-Dawley rats were fed hexachlorobenzene for 96 days prior to mating through 2 rounds of 

breeding (Kitchin et al. 1982). In both F1a and F1b pups, doses as low as 3 mg/kg/day decreased body 

weight and doses as low as 4 mg/kg/day decreased survival (Kitchin et al. 1982). 

The only animal developmental study in non-rodents used Rhesus monkey mothers (Bailey et al. 1980; 

Iatropoulos et al. 1978). Only one of three infant monkeys (between 21 and 118 days of age) nursing 

from mothers fed 64 mg/kg/day of hexachlorobenzene by daily gavage survived; the durations of dosing 

were 15 and 38 days (for the mortalities) and 60 days (for the survivor). Although the mothers were 

asymptomatic, both infant mortalities exhibited neurological effects (listlessness, lethargy, depression, 

and ataxia) and lung edema prior to death. Microscopic findings included mild hepatocellular 

hypertrophy in the infant that survived, and hepatic fatty changes, slight renal proximal tubule 

vacuolation, and mild cerebral gliosis in one or both infants that died. 

Two chronic-duration reproductive studies in Sprague-Dawley rats observed signs of developmental 

toxicity (Arnold et al. 1985; Grant et al. 1977). In one study male and female rats were fed hexachlorobenzene from 3 months prior to mating through weaning, and pups were continued on the same diet for 

their entire lifetime. Only the highest dose tested, 2 mg/kg/day, decreased pup survival. When examined 

as adults (week 130), peribiliary lymphocytosis and fibrosis were observed in F1 males at the lowest dose 

tested, 0.016 mg/kg/day, and hepatic basophilic chromogenesis was seen at $0.4 mg/kg/day (Arnold et al. 

1985). A four-generation assay found increased liver weight and hepatic aniline hydroxylase activity at 
HEXACHLOROBENZENE 111 

3. HEALTH EFFECTS 

2 mg/kg/day (F1a and F3a animals), consistently decreased pup weight at $4 mg/kg/day (all pup 

generations), and decreased pup viability at 8 mg/kg/day (F1a and F1b animals) (Grant et al. 1977). The 

LOAEL of 0.016 mg/kg/day from the lower-dose study (Arnold et al. 1985) has been used to calculate a 

chronic oral MRL of 0.00005 mg/kg/day as described in the footnote to Table 3-2 and in Appendix A. 

The highest NOAEL values and all reliable LOAEL values for developmental effects in each species and 

duration category are presented in Table 3-2 and plotted in Figure 3-2. 

3.2.2.7 Cancer 

The Department of Human Health Services (DHHS) considers the evidence for the carcinogenicity of 

hexachlorobenzene in experimental animals sufficient, and this chemical is reasonably anticipated to be a 

carcinogen in humans (NTP 2001). A cancer assessment for hexachlorobenzene is available on 

Integrated Risk Information System (IRIS 2001) in which the chemical is assigned to U.S. EPA cancer 

weight-of-evidence Group B2, probable human carcinogen, on the basis that oral administration of hexachlorobenzene has been shown to induce tumors in the liver, thyroid, and kidney in three rodent species. 

IRIS (2001) presents an oral slope factor of 1.6 per (mg/kg)/day and inhalation unit risk of 0.00046 per 

(μg/m3

) for hexachlorobenzene based on hepatocellular carcinoma in orally exposed female rats (Ertürk et 

al. 1986). This EPA assessment is currently undergoing re-evaluation. 

The available epidemiology reports taken together do not support an association between hexachlorobenzene exposure and increased cancer incidence, but their limitations (including small study sizes, 

similar tissue hexachlorobenzene levels between cancer and control groups, and potentially confounding 

effects of other organochlorines) preclude considering them evidence of noncarcinogenicity. Casecontrol studies (ranging in size from 20 cancer cases and 20 controls to 304 cases and 186 controls) 

investigating organochlorine levels in serum or breast tissue samples surgically removed from groups of 

patients either with breast cancer or with benign breast tumors were unable to detect statistically 

significant differences in hexachlorobenzene levels (Dorgan et al. 1999; Falck et al. 1992; Guttes et al. 

1998; Liljegren et al. 1998; Mendonca et al. 1999; Moysich et al. 1998; Mussalo-Rauhammaa et al. 1990; 

Zheng et al. 1999). The exception was a prospective study by Dewailly et al. (1994) in which serum and 

adipose organochlorine levels were measured in 41 Canadian women (ages 40-69) undergoing breast 

biopsy. Mammary adenocarcinoma was diagnosed in 20 women and benign breast tumors in 17 others; 

the former group had statistically significantly higher serum (but not adipose) levels of hexachlorobenzene compared to the latter group. No association of blood hexachlorobenzene levels with cancer 
HEXACHLOROBENZENE 112 

3. HEALTH EFFECTS 

incidence was observed in a study of 108 pancreatic cancer patients in the San Francisco Bay area 

(Hoppin et al. 2000), or in a study of 154 endometrial cancer patients in Sweden (Weiderpass et al. 2000). 

Additionally, studies analyzing hexachlorobenzene in the bone marrow of 13 German leukemia patients 

(Scheele et al. 1996) and in the adipose of 4 male Swedish patients with Ewing’s sarcoma of the bone 

(Hardell et al. 1997) did not find any statistically significant increases compared to controls. No evidence 

of cancer was reported in a 25-year follow-up study of 161 people (Peters et al. 1982) or a 20–30-year 

follow-up study of 204 people (Cripps et al. 1984) who consumed hexachlorobenzene-contaminated grain 

in Turkey from 1955 to 1959. However, these two studies did not examine patients for internal cancer 

and were not designed to detect increases in cancer incidence. 

Those follow-up studies (Cripps et al. 1984; Peters et al. 1982) did detect porphyria in some adults 

(17/204 and 33/161, respectively) who had been exposed as children to hexachlorobenzene (see 

Section 3.2.2.2 Hepatic Effects). This is relevant to cancer formation in humans because other 

epidemiology studies (unrelated to hexachlorobenzene) have found statistically significant associations 

between porphyria and increased risk of liver cancer. Fracanzani et al. (2001) reported that the presence 

of porphyria conferred a 5-fold increased risk of liver cancer. Linet et al. (1999) reported that porphyria 

cutanea tarda and acute intermittent porphyria were associated, respectively, with 20- and 70-fold 

increases in liver cancer and 3-fold increases in lung cancer. However, porphyria and liver cancer in the 

general population share common etiologies, so the association may not be casual (Axelson 1986; Salata 

et al. 1985; Topi et al. 1980; Waddington 1972). 

Oral exposure of rats, mice, and hamsters to hexachlorobenzene has induced tumors in the liver (“liver 

cell” tumor, hepatocellular carcinoma, hepatoma, hemangiohepatoma, hemangioendothelioma, bile duct 

tumor) (see below). Individual studies have also reported statistically significant increases in the 

incidences of kidney (renal cell adenoma), thyroid (alveolar adenoma), parathyroid (adenoma), and 

adrenal gland (pheochromocytoma) tumors, as well as the induction of lymphosarcoma (non-Hodgkin’s 

lymphoma) (Arnold et al. 1985; Ertürk et al. 1986). Female rats (Ertürk et al. 1986; Pereira et al. 1982; 

Smith and Cabral 1980) and mice (Cabral et al. 1979) appear to be more susceptible than males to the 

hepatocarcinogenic effects of hexachlorobenzene and limited evidence suggests that males are more 

susceptible to renal cancer (Ertürk et al. 1986). The cause of these gender specificities are unclear. 

Chronic oral exposure to hexachlorobenzene induces liver tumors in rats, with female rats appearing more 

susceptible than males. Ertürk et al. (1986) fed Sprague-Dawley rats 4 or 8 mg/kg/day of hexachlorobenzene for up to 2 years, with 9 interim sacrifices. Hexachlorobenzene induced statistically significant 
HEXACHLOROBENZENE 113 

3. HEALTH EFFECTS 

increases in the incidences of hepatoma, hepatocarcinoma, and renal carcinoma in both genders. In the 

liver, degenerative lesions were seen after 2–3 weeks, preneoplastic changes were detected after 200 days, 

and hepatocarcinomas were detected beginning at 300 days. Hepatomas, hemangiohepatomas, and 

hepatocellular carcinomas were significantly more common in females than in males. Bile duct adenomas 

(statistically significant increased incidence) and bile duct adenocarcinomas (not significant) were seen 

only in treated females. In contrast, renal adenomas and renal cell carcinomas were more frequent in 

males. Smith and Cabral (1980) detected liver cell tumors in all (14/14) female Agus rats and in a 

majority of female Wistar rats (4/6) fed hexachlorobenzene for 90 weeks at doses of 6–8 and 

5 mg/kg/day, respectively. Moreover, in those treated Wistar rats without tumors (2/6), evidence of 

preneoplastic changes (hepatocellular hypertrophy) was observed. Similar results were seen in a 

subsequent study (Smith et al. 1985). In Fischer 344/N rats fed 10 mg/kg/day of hexachlorobenzene for 

90 weeks, all surviving females had multiple liver tumors and at least 50% exhibited hepatocellular 

carcinomas. In contrast, only 16% of males exhibited liver tumors, which were smaller and limited to one 

per animal. Liver tumors stained heavily for gamma-glutamyl transpeptidase. In a tumor promotion 

study, feeding of 10 mg/kg/day of hexachlorobenzene for 45 days induced significantly more preneoplastic changes (hepatic foci positive for gamma-glutamyl transpeptidase) in the initiated livers of female 

Sprague-Dawley rats than of males (Pereira et al. 1982). In female (but not male) F1 Sprague-Dawley rats 

with lifetime exposure to 2 mg/kg/day of hexachlorobenzene, the incidence of liver neoplastic nodules 

was significantly increased compared to controls (Arnold et al. 1985). 

Hexachlorobenzene also induced liver tumors in mice and hamsters, but gender specificity was apparent 

only in mice. Syrian golden hamsters were fed 0, 4, 8, and 16 mg/kg/day of hexachlorobenzene in the 

diet for life (Cabral et al. 1977). Statistically significant increases in incidence were seen for hepatomas 

and hemangioendotheliomas (liver and spleen) in all treatment groups, with a slightly higher incidence in 

males than in females. Thyroid tumors were not seen in control groups, but were found in all treatment 

groups except low dose males; however, the increased incidence was significant only for males at 

16 mg/kg/day. Outbred Swiss mice were fed 6, 12, or 25 mg/kg/day of hexachlorobenzene for up to 

120 weeks (Cabral et al. 1979). Statistically significant increases in the incidence of liver cell tumors 

were seen at 12 and 25 (but not 6) mg/kg/day for both genders. Liver tumor incidence was significantly 

more common in females than in males treated with 25 mg/kg/day. Tumor multiplicity and size increased 

with increasing dose, while the latency period decreased. 

A brief paper by Ertürk et al. (1986) reported the induction of liver and other tumor types in SpragueDawley rats, Swiss mice, and Syrian golden hamsters after only 90 days of feeding the animals 20 or 
HEXACHLOROBENZENE 114 

3. HEALTH EFFECTS 

40 mg/kg/day of hexachlorobenzene. However, the report is limited by its lack of methodology and 

quantitative results. Liver effects included hepatomas, metaplasia, and stromal activation. Some support 

for the findings of a rapid onset of liver tumor formation is found in intermediate oral dosing (10 days to 

24 weeks) experiments with hexachlorobenzene that induced hepatocellular hypertrophy in rats (Den 

Besten et al. 1993; Smith et al. 1985), mice (Shirai et al. 1978), pigs (Den Tonkelaar et al. 1978), dogs 

(Sundlof et al. 1981), and Rhesus monkeys (Iatropoulos et al. 1976; Knauf and Hobson 1979). However, 

other intermediate-duration studies (13 weeks to 1 year) in rats did not detect neoplasia or preneoplastic 

effects in the liver or other organs (Goldstein et al. 1978; Kimbrough and Linder 1974; Kuiper-Goodman 

et al. 1977; Mollenhauer et al. 1975; Smith et al. 1979, 1985, 1986a). 

The mode-of-action for the hepatocarcinogenicity of hexachlorobenzene has been investigated. Multiple 

studies have indicated that non-heme iron potentiates the hepatocarcinogenic effects of hexachlorobenzene (Adjarov 1990; Elder and Urquhart 1986; Hahn et al. 1988; Smith and Francis 1983; Smith et al. 

1989, 1993; Vincent et al. 1989). Two studies in rats provided limited evidence that hexachlorobenzene 

is a promoter but not an initiator of liver cancer. As noted above, treatment of intact Sprague-Dawley rats 

with 10 mg/kg/day of hexachlorobenzene for 45 days did not induce hepatic foci positive for gammaglutamyl transpeptidase, but foci were induced following liver initiation by partial hepatectomy with and 

without diethylnitrosomaine (Pereira et al. 1982). This finding suggests that hexachlorobenzene may act 

as an promoter at doses insufficient to initiate tumors. No evidence of altered foci were detected in male 

Fisher 344 rats pretreated with partial hepatectomy and treated with a single gavage dose of 5,000 mg/kg 

hexachlorobenzene followed by liver tumor promotion with carbon tetrachloride and cholic acid for 

12 weeks (Tsuda et al. 1993). 

Reported incidences of other tumor types have also increased following oral exposure to hexachlorobenzene. A group of five adult female Rhesus monkeys were dosed by gavage with hexachlorobenzene 

in methylcellulose daily for 60 days (Iatropoulos et al. 1976). Single monkeys received 0, 8, 32, or 

64 mg/kg and two monkeys received 128 mg/kg. In one of the high dose monkeys (but not in the others), 

a benign mammary fibroadenoma was detected. This evidence is inconclusive, because of the low 

statistical power of the study. The other monkey given 128 mg/kg exhibited moderate adrenal medullary 

hyperplasia and the monkey given 64 mg/kg exhibited slight hyperplasia of the adrenal zona fasciculata; 

these findings support observations made in rats. In F1 Sprague-Dawley rats exposed to 2 mg/kg/day for 

their entire lives (including in utero exposure and approximately 130 weeks ex utero, as part of a 

2-generation reproductive study), statistically significant increases in adrenal pheochromocytomas were 

seen in both males and females, while a significantly increased incidence of parathyroid adenomas was 
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3. HEALTH EFFECTS 

observed only in females (Arnold et al. 1985). The biological significance of the adrenal effect is also 

supported by observations of adrenal gland cortical hyperplasia in female Wistar rats at doses as low as 

9.5 mg/kg/day for 13 weeks (Den Besten et al. 1993) and in male and female Sherman rats exposed to 

hexachlorobenzene in the diet at doses at low as 5 mg/kg/day for 4 months (Kimbrough and Linder 1974). 

No other studies have reported any parathyroid histopathology caused by oral exposure to hexachlorobenzene, but parathyroid effects (changes in hormone levels) have been observed in male Fischer 344 rats 

(Andrews et al. 1988, 1989, 1990). The observations made during the 90-day study in which SpragueDawley rats, Swiss mice, and Syrian golden hamsters were fed 20 or 40 mg/kg/day of hexachlorobenzene have not been confirmed (Ertürk et al. 1986). Renal damage with metaplastic regenerative 

changes were seen in treated animals; renal tumors were “most frequent in rats” and “more frequent in 

males.” Lymphosarcomas (detected in the thymus, spleen, and lymph nodes) were apparently common in 

all treatment groups, with frequent lymphohematopoietic hyperplasia and lymphocytic infiltrations. 

These lesions occurred 2–4-fold (depending on organ) more frequently in female mice than in male mice. 

The CEL (i.e., lowest dose that produced a tumorigenic response for each species), the duration category 

of exposure to hexachlorobenzene, and the estimated upper-bound risk levels from 10-4 to 10-7 are shown 

in Table 3-2 and plotted in Figure 3-2. 

3.2.3 Dermal Exposure 

No studies were located regarding the following health effects in humans or animals following dermal 

exposure to hexachlorobenzene: 

3.2.3.1 Death 

3.2.3.2 Systemic Effects 

3.2.3.3 Immunological and Lymphoreticular Effects 

3.2.3.4 Neurological Effects 

3.2.3.5 Reproductive Effects 

3.2.3.6 Developmental Effects 
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3. HEALTH EFFECTS 

3.2.3.7 Cancer 

3.3 GENOTOXICITY 

An increased incidence of micronuclei was observed in the peripheral lymphocytes of 41 chemical 

workers in San Paulo, Brazil, who had been exposed to a mixture of chlorinated solvents that included 

hexachlorobenzene, as well as carbon tetrachloride and perchlorethylene (da Silva Augusto et al. 1997). 

The usefulness of this study is limited by the confounding effect of multiple chemical exposure. 

No studies were located regarding genotoxic effects in animals following inhalation exposure to hexachlorobenzene. 

No studies were located regarding the genotoxic effects of hexachlorobenzene in humans following oral 

exposure. 

In vivo studies in rats revealed the lack of significant genotoxic activity in mammals following oral 

exposures to hexachlorobenzene. Negative results were observed in two dominant lethal mutation assays 

that orally exposed rats at doses ranging from 60 to 221 mg/kg (Khera 1974a; Simon et al. 1979). No 

evidence of genotoxicity was observed in mouse liver, lung, kidney, spleen, or bone marrow after oral 

dosing (Sasaki et al. 1997). An oral exposure study to test the DNA induction potential of hexachlorobenzene in Wistar rats provided equivocal evidence that hexachlorobenzene reacts directly with DNA 

(Gopalaswamy and Nair 1992). Male rats were untreated or pretreated with phenobarbital (0.1% sodium 

phenobarbital in drinking water for 2 weeks) and then administered 25 mg/kg (specific activity 

14.0 mCi/mmole) hexachlorobenzene in 0.1 mL refined peanut oil for 24 hours. The animals were 

sacrificed and DNA obtained from liver extracts. Upon analysis, the DNA was found to be bound by 

hexachlorobenzene (2.23±0.27 pmoles/mg DNA for phenobarbital untreated animals and 

3.56±0.18 pmoles/mg DNA for phenobarbital pretreated animals). The comparative values for lindane in 

the same study were 5.82±0.31 and 6.90±0.14 pmoles/mg DNA , respectively. No hexachlorobenzene 

untreated control values were provided in the study report. It is notable, however, that there is evidence 

(Jackson et al. 1993) that phenobarbital is mutagenic in vitro in several test systems. Other studies have 

likewise failed to observed gene mutations or unscheduled DNA repair in microbial assays 

(Gopalaswamy and Nair 1992; Haworth et al. 1983; Siekel et al. 1991). 
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No studies were located regarding genotoxic effects in humans or animals following dermal exposure to 

hexachlorobenzene. 

Hexachlorobenzene did not produce chromosomal aberrations in human peripheral lymphocytes in vitro 

(Siekel et al. 1991). However, hexachlorobenzene produced weak positive results in assays for DNA 

fragmentation and micronuclei formation in primary cultures of human hepatocytes (Canonero et al. 

1997). Treatment with hexachlorobenzene induced only minimal formation of DNA adducts in cultured 

human Hep G2 hepatoma cells (Dubois et al. 1997). 

The micronucleus assay, but not the DNA fragmentation assay, was positive in cultured rat hepatocytes 

(Canonero et al. 1997). The researchers concluded that hexachlorobenzene is a weak genotoxic 

carcinogen and that negative responses in standard genotoxicity assays were due to limitations in the 

ability of exogenous metabolic activation systems to duplicate the complex interactions of the intact liver 

cell. Hexachlorobenzene was also positive in an assay for replicative DNA synthesis in mouse 

hepatocytes (Miyagawa et al. 1995). Treatment with hexachlorobenzene induced only minimal formation 

of DNA adducts in fetal hepatocytes from rats and quail (Dubois et al. 1997). 

Hexachlorobenzene tested negative or ambiguous in reverse mutation assays in S. typhimurium 

(Gopalaswamy and Aiyar 1986; Gopalaswamy and Nair 1992; Haworth et al. 1983; Siekel et al. 1991) 

and E. coli (Siekel et al. 1991) with and without metabolic activation, although an assay for reverse 

mutation in the yeast Saccharomyces cerevisiae was positive (Guerzoni et al. 1976). Hexachlorobenzene 

also tested negative in a DNA repair assay in E. coli (Siekel et al. 1991). 

Data from a study in male Long Evans rats suggested that the metabolism of hexachlorobenzene to pentachlorobenzene and other more polar metabolites proceed either through a free-radical mechanism or by 

initial formation of an arene oxide. These reactive intermediates may form covalent bonds with cellular 

constituents (such as protein amino acids or DNA nucleic acids) leading to irreversible cell damage (Lui 

and Sweeney 1975). Several other studies have also found evidence of binding to cellular proteins by 

reactive electrophilic metabolites of hexachlorobenzene formed by cytochrome P-450 system 

(Gopalaswamy and Nair 1992; Rajamanickam and Padmanaban 1974; van Ommen et al. 1985). 

Key in vivo genotoxicity studies are presented in Table 3-3 and in vitro genotoxicity studies are presented 

in Table 3-4. 
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3. HEALTH EFFECTS 

3.4 TOXICOKINETICS 

In humans, inhalation accounts for an unknown, but probably low, amount of exposure due to the low 

vapor pressure of hexachlorobenzene. Current information indicates that human absorption of inhaled 

hexachlorobenzene is poor; approximately two orders of magnitude less than the exposure estimate for 

the oral route (Burns et al. 1974; Burton and Bennett 1987; Currier et al. 1980). Other data of absorption 

following inhalation exposure come from studies of occupational exposures (Queiroz et al. 1997, 1998a, 

1998b; Richter et al. 1994; Selden et al. 1997) and citizens of Flix, Spain, who have been exposed to 

airborne hexachlorobenzene from an organochlorine factory (Grimalt et al. 1994; Herrero et al. 1999; Sala 

et al. 1999b; To-Figueras et al. 1997). Based on information from an epidemic resulting from ingestion 

of hexachlorobenzene-contaminated bread in Turkey, ingested hexachlorobenzene is moderately absorbed 

from the gastrointestinal tract (Albro and Thomas 1974; Cam and Nigogosyan 1963; Gocmen et al. 1989; 

Peters et al. 1982). However, most of the hexachlorobenzene body burden in the U.S. population derives 

from dietary intake of fatty foods (Burton and Bennett 1987). Schlummer et al. (1998) estimated that 

85.4% of ingested hexachlorobenzene will be absorbed when the blood contains no hexachlorobenzene, 

and that this percentage will be reduced by 0.2% for each ng of hexachlorobenzene per g lipid in blood, 

and hypothesized a “fat-flush” theory of hexachlorobenzene absorption: temporary increases in lipid 

content in the gut dilute hexachlorobenzene concentrations and increase the diffusion gradient from the 

gut into the lymph and blood. Data from animal studies indicate that the gastrointestinal absorption of 

hexachlorobenzene is quite variable, depending upon the solvent vehicle used for administration, ranging 

from 6% when administered in aqueous solution to 82% when administered with squalene in cottonseed 

oil (Albro and Thomas 1974), olive oil (Freeman et al. 1989; 
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3. HEALTH EFFECTS 

Table 3-3. Genotoxicity of Hexachlorobenzene In Vivo 

Species Exposure 

End point (test system) route Results Reference 

Mammalian systems:

 Dominant lethals Rat Oral – Khera 1974a 

Rat Oral	 – Simon et al. 1979

 DNA binding Rat (Wistar) Oral ±	 Gopalaswamy and Nair 

1992 

– = negative result; ± = weakly positive; DNA = deoxyribonucleic acid 
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3. HEALTH EFFECTS 

Table 3-4. Genotoxicity of Hexachlorobenzene In Vitro 

Results 

With Without 

Species (test system) End point	 Reference activation activation 

Prokaryotic organisms: 

Reverse mutation 

Salmonella typhimurium Gene mutation – – Haworth et al. 1983 

S. typhimurium	 Gene mutation ± – Gopalaswamy and 

Nair 1992 

S. typhimurium Gene mutation – –	 Siekel et al. 1991 

Escherichia coli Gene mutation – – Siekel et al. 1991 

DNA Repair Assays 

E. coli	 DNA repair – – Siekel et al. 1991 

Eukaryotic cells: 

Human peripheral blood Chromosomal – – Siekel et al. 1991 

lymphocytes aberration 

Mammalian system: 

Rat (Wistar) DNA binding ± – Gopalaswamy and 

Nair 1992 

– = negative result; ± = weakly positive; DNA = deoxyribonucleic acid 
HEXACHLOROBENZENE 121 

3. HEALTH EFFECTS 

Goldey et al. 1990; Knauf and Hobson 1979; Koss and Koransky 1975; Mehendale et al. 1975; Sundlof et 

al. 1982; Villeneuve and Hierlihy 1975), or peanut oil (Ingebrigtsen and Nafstad 1983; Ingebrigtsen et al. 

1981). The lymphatic system has been shown to play an important role in the gastrointestinal uptake of 

hexachlorobenzene in animals (Iatropoulos et al. 1975). Although no empirical data are available on the 

dermal absorption of hexachlorobenzene in humans, data from a rat study were used to develop a 

compartment model for application to a 70-kg worker. Using these data, the dermal absorption constant 

for hexachlorobenzene was calculated as 1.4x10-3 per hour (Koizumi 1991). 

Information on distribution in people following inhalation exposure to hexachlorobenzene is limited 

(Ataniyazova et al. 2001; Sala et al. 2001b) and no information is available on the distribution of inhaled 

hexachlorobenzene in animals. However, orally absorbed hexachlorobenzene distributes widely in 

mammalian tissue, rapidly partitioning to blood, liver, breast milk, adipose tissue, endocrine organs, bone 

marrow, and ovarian follicular fluid (Ellenhorn and Barceloux 1988; Foster et al. 1995a; Ingebrigtsen 

1986; Ingebrigtsen and Nafstad 1983; Knauf and Hobson 1979; Wickstrom et al. 1983), and preferentially 

distributing to adipose tissue or organs with high fat content (Burton and Bennett 1987; Cohn et al. 1978; 

Koss and Koransky 1975; Lecavalier et al. 1994; Mehendale et al. 1975; Robinson et al. 1990; Teufel et 

al. 1990; van Raaij et al. 1993a; Verschueren 1983). Animal studies of oral dosing have showed that 

levels of hexachlorobenzene increase in a dose-dependent manner in all tissues up to 100 mg/kg/day 

(Foster et al. 1995a; Jarrell et al. 1993; Sundlof et al. 1982). Hexachlorobenzene body burden is readily 

transferred from pregnant mother to the fetus through the placenta in animals (Courtney and Andrews 

1985; Courtney et al. 1979; Cripps 1990; Goldey et al. 1990; Nakashima and Ikegari 2000; Nakashima et 

al. 1997, 1999; Villeneuve and Hierlihy 1975; Villeneuve et al. 1974a). Additionally, hexachlorobenzene 

is concentrated in the milk and can be transferred to the suckling neonate (Bailey et al. 1980; Cripps 

1990; Goldey et al. 1990; Nakashima and Ikegari 2000; Nakashima et al. 1997, 1999). Evidence from 

animal studies indicates that protein-poor diets may promote the preferential partitioning of ingested 

hexachlorobenzene to fatty tissue (Rodrigues et al. 1991). In a survey of the U.S. population, it was 

found that concentration of hexachlorobenzene tended to increase with increasing age, a testimony to the 

propensity of hexachlorobenzene to bioaccumulate in mammalian tissue (Robinson et al. 1990). In a 

group of 350 German children, blood hexachlorobenzene levels (and levels of other organochlorines) 

correlated strongly with the length of breast-feeding (Karmaus et al. 2001). Weak associations were seen 

between decreased blood hexachlorobenzene levels and increased child body mass index (above 

18 kg/m2

), and between increased hexachlorobenzene levels and both maternal age at birth (36–45-yearold group only) and late birth order (3rd or later, with spacing between children of at least 4 years). These 

data suggest that increased size may dilute hexachlorobenzene in the body, and that levels of hexachloro­
HEXACHLOROBENZENE 122 

3. HEALTH EFFECTS 

benzene in mothers may increase with age. No correlations were seen for mothers who smoked during 

pregnancy, or the age and gender of the child. 

Hexachlorobenzene is slowly metabolized in mammals, and the majority of hexachlorobenzene is 

excreted unchanged (in feces). Reductive dechlorination of hexachlorobenzene—catalyzed by enzymes 

located in the microsomal fraction of liver, lung, kidney, and intestine—appears to be an important 

pathway for the metabolism of hexachlorobenzene (Ingebrigtsen et al. 1986). It has been suggested that 

epoxide formation also occurs in this metabolism (Lui et al. 1976). PCP has been identified in human 

liver preparations incubated with hexachlorobenzene (Koss et al. 1986). In animals, hexachlorobenzene 

is slowly metabolized to PCP by the hepatic cytochrome P-450 system, conjugated with glutathione to 

yield pentachlorothiophenol, or reductively dechlorinated to form pentachlorobenzene. Other metabolites 

include less chlorinated benzenes, chlorophenols, S-conjugated phenols, and benzenes. PCP is 

subsequently converted to TCHQ (Hahn et al. 1988, 1989; Linko et al. 1986; Mehendale et al. 1975; 

Rozman et al. 1977a). The feces contain mostly unchanged parent compound, and about 1% pentachlorobenzene and traces of PCP after oral hexachlorobenzene exposure in mammals, while urinary excretion 

consists of mostly the metabolites, pentachlorobenzene, 2,4,5-trichlorophenol, N-acetyl-S(pentachlorophenyl)cysteine, mercaptotetrachlorothioanisole, and tetrachlorobenzene, 2,3,5,6-tetrachlorobenzene1,4-diol; and unchanged parent compound (Koss et al. 1978; Mehendale et al. 1975; Rizzardini and Smith 

1982; Rozman et al. 1978). Pentachlorothiophenol, PCP, methylthiopentachlorobenzene, 1,4-bis­

(methylthio)-2,3,5,6-tetrachlorobenzene, chlorophenols, S-conjugated phenols and benzenes, and less 

chlorinated benzenes have also been identified in the liver following oral exposure in animals (D'Amour 

and Charbonneau 1992; Engst et al. 1976a; Ingebrigtsen et al. 1981, 1986; Jansson and Bergman 1978; 

Koss et al. 1976, 1979; Lui and Sweeney 1975; Renner 1988; Richter et al. 1981; Stewart and Smith 

1986; To-Figueras et al. 1992; van Ommen et al. 1985, 1989; Yang et al. 1978). Sex differences in the 

metabolism of hexachlorobenzene in the adult animals have been reported. Urinary excretion of PCP, 

2,3,5,6-tetrachlorobenzene-1,4-diol, and pentachlorothiophenol was greater in females than in males in 

this study (Rizzardini and Smith 1982). 

To-Figueras et al. (2000) observed a high correlation between fecal and blood levels of hexachlorobenzene in 53 people highly exposed to airborne hexachlorobenzene. No information is available on the 

excretion of hexachlorobenzene following inhalation exposure in animals or following dermal exposure 

in humans or animals. In humans, ingested hexachlorobenzene is excreted in the urine mainly as its 

metabolites, PCP and pentachlorothiophenol (To-Figueras et al. 1992). In animals, the excretion of hexachlorobenzene appears to be quite variable, depending upon the solvent vehicle used (Albro and Thomas 
HEXACHLOROBENZENE 123 

3. HEALTH EFFECTS 

1974; Rozman et al. 1977a). Based on decreasing concentrations in the liver, the biological half-life of 

hexachlorobenzene has been estimated to be 8 days at the start of the elimination phase, 10 weeks after 

3 months, and 12 months after 1.5 years (Koss et al. 1983), suggesting differential release of hexachlorobenzene from tissue stores, perhaps as a function of lipophilicity. Ingested hexachlorobenzene is excreted 

predominantly in the feces, mainly as unchanged parent compound, and to a lesser extent in the urine, as 

its metabolites (PCP, pentachlorothiophenol, pentachlorobenzene) (Mehendale et al. 1975). 

Approximately 99% of unchanged ingested hexachlorobenzene was excreted in the feces; 50% of urinary 

excretion was PCP, 25% was pentachlorobenzene, and 25% was unchanged hexachlorobenzene in Rhesus 

monkeys treated with 0.03 mg/kg/day in the diet for 15 months (Rozman et al. 1977a). Based on animal 

studies, the urinary excretion of hexachlorobenzene exhibits sex- and age-specific differences; the 

excretion of pentachlorothiophenol increases with sexual maturity in female rats and slightly decreases in 

male rats (To-Figueras et al. 1991). Biliary excretion was not an important excretory pathway in rats 

given a single hexachlorobenzene dose of 10 mg/kg by gavage in peanut oil, accounting for <4% of the 

administered dose (Ingebrigtsen et al. 1981). 

3.4.1 Absorption 

3.4.1.1 Inhalation Exposure 

Limited data show that hexachlorobenzene can be absorbed through the respiratory tract in humans, 

although no information is available as to the rate and extent of respiratory tract absorption of hexachlorobenzene in either humans or animals. 

Spanish researchers have studied a population with long-term exposure to high levels of hexachlorobenzene in air (Grimalt et al. 1994; Herrero et al. 1999; Sala et al. 1999b; To-Figueras et al. 1997). The 

rural Spanish village of Flix contains an organochlorine factory that has been producing volatile 

chlorinated solvents for decades, and no other large industrial facilities. Following complaints of odor, 

approximately 40 air samples were collected in July and November of 1989 and May and October of 

1992 at diverse sites in the village. As a control, five air samples were collected in the city of Barcelona. 

Average air levels of hexachlorobenzene in Flix (35 ng/m3

) were over 100-fold higher than in Barcelona 

(0.3 ng/m3

), while other organochlorines were found at similar or lower concentrations in Flix than in 

Barcelona. Corresponding to the high air levels, it was found that residents of Flix had unusually high 

serum levels of hexachlorobenzene (mean of 39.8 ng/mL based on a total number of 604 tested) in 

comparison to populations in Barcelona (mean=4.13 ng/mL, n=100), the United States 
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3. HEALTH EFFECTS 

(mean=0.19 ng/mL, n=370), Croatia (mean=1.00, n=15), and Germany (mean=1.12, n=6). Serum levels 

of other organochlorines in Flix residents were much lower than hexachlorobenzene levels and did not 

differ from other populations. Among Flix residents, serum hexachlorobenzene levels were several fold 

higher in factory workers (mean=93.4 ng/mL, n=185) than other residents (mean=16.9 ng/mL, n=419). 

Factory workers were presumably exposed to much higher air levels of hexachlorobenzene than other 

village residents, and some may have had dermal exposure as well. 

It is noteworthy that mean serum hexachlorobenzene levels in Flix residents who did not work at the 

factory (and therefore, can be assumed to have had no direct dermal exposure to hexachlorobenzene) were 

still 4-fold higher than Barcelona residents. However, the difference was not entirely due to inhalation 

exposure. In addition to working at the factory, other variables associated with serum hexachlorobenzene 

levels were age and consumption of local fish. Among women (very few of whom had ever worked at the 

factory), the geometric mean serum hexachlorobenzene level was 14.9 ng/mL in those that did not eat 

local fish (176/180) and 18.2 ng/mL in those that did (only 4/180). Therefore, indirect exposure to hexachlorobenzene via consumption of contaminated fish may have contributed slightly to serum hexachlorobenzene levels in nonfactory worker Flix residents, but was not a major factor. This study shows that 

exposure to high levels of hexachlorobenzene in air leads to high levels of hexachlorobenzene in serum, 

and that a significant portion of hexachlorobenzene uptake in this situation can be attributed to inhalation 

and absorption across the respiratory tract. 

Studies of workers with occupational exposure to hexachlorobenzene, where exposure was probably 

primarily by inhalation, but may have involved dermal contact as well, also show increased serum levels 

of hexachlorobenzene in the exposed workers. Selden et al. (1997) found significantly higher serum 

hexachlorobenzene levels in 29 hazardous waste incineration workers (63 ng/g lipid) than in 60 matched 

controls (35 ng/g lipid). The exposed workers also had significantly increased serum hexachlorobenzene 

levels compared with their own historical samples given before the start of employment (0.40 ng/g plasma 

vs. 0.27 ng/g plasma). Airborne hexachlorobenzene levels in different locations in the plant ranged from 

0.066 to 11 ng/m3

. Queiroz et al. (1997, 1998a, 1998b) observed that each of the 51 workers on leave 

from a closed chemical plant had blood hexachlorobenzene levels >0.1 μg/dL (mean=4.4 μg/dL), while 

controls chosen from blood donors at the local blood bank to be similar in age and race to the exposed 

group all had blood hexachlorobenzene levels lower than the limit of detection (0.02 μg/dL). The plant 

produced carbon tetrachloride and tetrachloroethylene; hexachlorobenzene was generated as a byproduct 

of the production process as a solid residue. Richter et al. (1994) documented high serum hexachlorobenzene levels in workers exposed to 2.1–10.8 mg/m3

 of hexachlorobenzene in air, which persisted even 
HEXACHLOROBENZENE 125 

3. HEALTH EFFECTS 

after air concentrations of hexachlorobenzene were reduced to 0.012–0.022 mg/m3. Although dermal 

exposure cannot be ruled out in these studies, uptake of hexachlorobenzene across the respiratory tract is 

likely to have contributed significantly to hexachlorobenzene body burden in all of these studies. 

No studies were located regarding inhalation exposure to hexachlorobenzene in animals. 

3.4.1.2 Oral Exposure 

Widespread occurrence of porphyria cutanea tarda in southeast Anatolia in Turkey in the late 1950s was 

shown to be due to ingestion of bread made from grain that had been treated with hexachlorobenzene 

(Cam and Nigogosyan 1963). The ingested dose of hexachlorobenzene was estimated to be in the range 

of 0.05–0.2 g/day, or 0.7–2.9 mg/kg/day for a 70-kg person. The occurrence of systemic health effects 

following ingestion of hexachlorobenzene demonstrates that this chemical can be absorbed via the 

gastrointestinal tract in humans, and in amounts sufficient to produce serious health effects. Follow-up 

studies conducted between 1977 and 1981 found that hexachlorobenzene levels in 56 samples of human 

milk obtained from porphyric mothers averaged 0.51 ppm (standard deviation [SD]=0.75 ppm, highest 

value=2.8 ppm), while levels in women from families without porphyria or outside the affected area had 

an average level of 0.07 ppm (SD=0.07 ppm) (Peters et al. 1982). Therefore, even 20 years after 

exposure, there was a large difference in breast milk hexachlorobenzene concentrations between people 

from families that had consumed the contaminated grain and those that did not. 

One study was located in which gastrointestinal absorption of hexachlorobenzene in humans was 

quantified. Schlummer et al. (1998) used a mass-balance approach to estimate absorption of hexachlorobenzene ingested at low concentrations in the diet in seven volunteers (four males and three females) 

ranging in age from 24 to 81 years. Hexachlorobenzene was measured in the food (uniform meals of 

varying portion sizes) ingested by volunteers over a 3-day period (using duplicate portions) and in the 

corresponding feces (first and last meals identified using iron capsules to produce black feces). Similar 

experiments were then conducted in which the volunteers chose their own foods. Whole blood samples 

were collected 3 weeks after the last mass balance experiment. Percent net absorption was calculated as 

the difference between ingested and excreted hexachlorobenzene, divided by the ingested amount. When 

fed a standardized meal, the percent absorption was a relatively uniform 70–82% in the four young adults 

tested (one female and three males ranging in age from 24 to 36 years). It decreased to 1% in a 53-yearold male volunteer and further to -56 and -210% in 76- and 81-year-old female volunteers, respectively. 
HEXACHLOROBENZENE 126 

3. HEALTH EFFECTS 

The negative values in the elderly volunteers indicate net excretion, rather than absorption, in these 

individuals. Similar results were reportedly obtained when volunteers chose their own meals. 

Blood levels of hexachlorobenzene (expressed as ng/g blood lipid) also varied with age, ranging from 

65 to 82 in the young adult volunteers and increasing to 230, 680, and 1,420 in the 53-, 76-, and 81-yearold volunteers, respectively (Schlummer et al. 1998). The researchers noted a trend for decreasing 

absorption with increasing blood levels in the volunteers, and a linear regression was fit to these data. 

The calculated regression equation was: 

percent absorption hexachlorobenzene = 0.8538–0.0021 x (ng hexachlorobenzene/g blood lipid) 

This equation suggests that 85.4% of ingested hexachlorobenzene will be absorbed when the blood 

contains no hexachlorobenzene, and that this percentage will be reduced by 0.2% for each ng of hexachlorobenzene per g lipid in blood. This regression equation provided an excellent fit to the data 

(R2=0.98), leading to the conclusion that concentration of hexachlorobenzene in the blood is a major 

factor influencing absorption of hexachlorobenzene from the gut. Although absorption of lipophilic 

compounds such as hexachlorobenzene is thought to occur by simple passive diffusion, analysis of the 

entire data set in this study (17 polychlorinated dibenzo-p-dioxins/furans [PCDD/F] congeners and 

10 PCBs in addition to hexachlorobenzene) showed that absorption of these compounds could not be 

explained by simple gradients between lipid-based food and blood concentrations or between gut contents 

and gut tissue. The researchers offered a “fat-flush” hypothesis of absorption for hexachlorobenzene and 

related compounds whereby lipid absorbed into gut tissue would temporarily increase lipid content of the 

tissue, diluting contaminant concentration in the gut tissue (on a lipid basis) and increasing the diffusion 

gradient for the contaminant from the gut contents, before being transported from the gut tissue to the 

lymph and blood. 

Gastrointestinal absorption of hexachlorobenzene has been well studied in laboratory animals. 

Ingebrigtsen and Nafstad (1983) demonstrated that gastrointestinal absorption of hexachlorobenzene in an 

oil vehicle is rapid. Male Wistar rats were given a single gavage dose of 0.4 mg/kg of radiolabeled hexachlorobenzene dissolved in peanut oil and examined by whole body autoradiography at various time 

intervals starting 2 hours after dosing. A considerable amount of radiolabeled material was absorbed and 

distributed throughout the body within 2 hours of dosing, and peak levels were reached within 24 hours. 

Rapid absorption was also shown in beagle dogs that had hexachlorobenzene levels in blood monitored 

during and after 7 days of daily oral dosing with 10 or 100 mg/kg of hexachlorobenzene in corn oil by 
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capsule (Sundlof et al. 1982). Peak blood concentrations occurred 3 hours after dosing in low-dose dogs, 

and after a somewhat longer (unspecified) interval in high-dose dogs. Blood concentrations continued to 

increase over several days after the last dose was administered in both groups, possibly due to continued 

absorption from the intestines during that time. This finding suggests that while the bulk of an oral dose 

of hexachlorobenzene in oil is absorbed rapidly in a few hours, absorption of residual quantities can 

continue for a period of days. 

Hexachlorobenzene administered by gavage in aqueous methylcellulose suspension is also absorbed from 

the gut within a few hours (Iatropoulos et al. 1975). Sprague-Dawley rats given single gavage doses of 

0.15 mg (approximately 0.6 mg/kg) of radiolabeled hexachlorobenzene in 1% methylcellulose solution 

were sacrificed at intervals between 1 and 48 hours after dosing for determination of tissue radioactivity 

levels. The ingested material was absorbed by the walls of the stomach and duodenum within 1 hour of 

dosing, and by the jejunum and ileum within 3 hours. Peak levels in the duodenum and jejuno-ileum 

occurred 3 hours after dosing. The majority of ingested hexachlorobenzene was absorbed from these 

regions of the small intestine by the lymphatic system and deposited in the fat, bypassing portal 

circulation to the liver, systemic circulation, and the excretory organs. 

Although absorption of hexachlorobenzene from aqueous suspension occurred in a similar time frame as 

absorption from an oil vehicle in the studies described above, other studies have shown that the extent of 

absorption from aqueous suspension is much less. Koss and Koransky (1975) measured absorption of 

radiolabeled hexachlorobenzene in female Wistar rats following gavage administration of the chemical in 

olive oil at doses of 20, 60, and 180 mg/kg, and in aqueous suspension (6% gum arabic in water) at doses 

of 16, 120, and 970 mg/kg. Two days after dosing in olive oil, 73–88% of the administered radioactivity 

was recovered in the body, while 1% was recovered in the gut contents, 18–26% in the feces, and 

0.4–0.6% in the urine in the different dose groups. This finding suggests oral absorption of about 80% of 

ingested hexachlorobenzene, regardless of dose, when administered in oil. When administered in aqueous 

suspension, however, absorption was much lower and appeared to depend on dose. At the low dose of 

16 mg/kg, roughly 20% of administered radioactivity was recovered in body tissues 3 days after dosing in 

aqueous suspension, compared with 0.4% in the gut contents, 74% in the feces, and 0.4% in the urine. At 

the higher doses, only 2–5% of the administered hexachlorobenzene was absorbed from the aqueous 

suspension. 

Other studies determined absorption of hexachlorobenzene from oil vehicles to be similar to that reported 

by Koss and Koransky (1975). Albro and Thomas (1974) gave male CD rats single gavage doses of 12 or 
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30 mg/kg of hexachlorobenzene in cottonseed oil. They found that after 96 hours, 72% (high dose) to 

82% (low dose) of the dose had not been excreted in the feces. No hexachlorobenzene was detected in 

the bile or urine, and only about 3% of the dose was present in the intestinal tissue and contents (primarily 

the former), and an associated in vitro experiment showed that fecal bacteria do not metabolize hexachlorobenzene; this suggests that the “removed” 72–82% had been absorbed into the body. Ingebrigtsen 

et al. (1981) used bile duct cannulated rats to monitor biliary excretion of radiolabeled hexachlorobenzene 

after gavage dosing with 10 mg/kg in peanut oil. A total of 3.6% of the administered radioactivity was 

recovered in the bile within 48 hours, while bile flow remained steady, showing that biliary excretion is 

only a minor pathway for hexachlorobenzene. After 96 hours, approximately 25% of the administered 

radioactivity was recovered in the feces and 2% in the urine. These data again suggest oral absorption of 

somewhere near 80% of the ingested dose for hexachlorobenzene administered in oil. 

The studies described above showed that absorption of hexachlorobenzene in the gut is much more 

extensive from oil vehicles than from aqueous vehicles. Zabick and Schemmel (1980) demonstrated that 

a high fat diet also enhances absorption of hexachlorobenzene, in comparison to a high carbohydrate diet. 

Groups of 6 female Osbourne-Mendel rats were fed either a high fat diet or one of two high carbohydrate 

diets (one using corn starch, the other using sucrose) supplemented with 32 mg/kg/day of hexachlorobenzene for 6, 12, or 18 days. The high fat diet resulted in higher carcass fat content (data not shown; 

cited to Shier and Schemmel 1975) and greatly increased concentrations of hexachlorobenzene in 

perirenal fat, liver, and gastrocnemius. At the same time, concentrations of hexachlorobenzene in the 

feces were much lower with the high fat diet, suggesting that the high fat diet facilitated absorption of 

hexachlorobenzene from the gut, thereby leading to the increase in tissue levels. The data from all of 

these studies showing enhanced absorption of hexachlorobenzene when administered in oil or a high fat 

diet are consistent with the “fat-flush” hypothesis for hexachlorobenzene absorption proposed by 

Schlummer et al. (1998) based on the human data. 

3.4.1.3 Dermal Exposure 

No studies were located regarding absorption in humans following dermal exposure to hexachlorobenzene. 

Evidence from rats suggests that hexachlorobenzene can be absorbed across the skin. Using radiolabelled 

hexachlorobenzene, Koizumi (1991) conducted a mass-balance study of dermal absorption in male Fisher 

344 rats. A dermal dose of approximately 2.5 mg/kg of 14C-hexachlorobenzene dissolved in 
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tetrachloroethylene was applied to a 4 cm2 clipped area on the back under occlusion. The rats, 3 per 

group, were transferred to metabolic cages and sacrificed after 6, 24, or 72 hours. Cumulative absorption 

of hexachlorobenzene (the sum recovered from the urine, feces, liver, carcass, skin not directly 

contaminated, and subcutaneous tissue) increased with duration of exposure from 1.05% of the applied 

dose after 6 hours to 2.67% after 24 hours and 9.71% after 72 hours. A one-compartment linear 

pharmacokinetic model was used to calculate an absorption constant of 1.40x10-3 per hour. 

A modeling exercise conducted by Koizumi (1991) suggests that dermal absorption of hexachlorobenzene 

can contribute significantly to body burden in exposed workers. Assuming the rate constant in rats 

applies to man and a biological half-life ranging from 100 to 730 days, a three-compartment linear 

pharmacokinetic model developed based on the rat data and scaled up to a 70-kg man showed that hexachlorobenzene blood levels will increase with duration of exposure and that dermal doses as low as 

2.56–18.2 mg (which could result from contamination small enough to go unnoticed) could, over a period 

of years, lead to hexachlorobenzene blood levels in the vicinity of 200 ppb, regarded by some researchers 

(Currier et al. 1980) as the upper safe limit in humans. 

Koizumi (1991) also collected data showing that washing can decrease the absorption of dermallycontacted hexachlorobenzene by a significant degree. In the rats, washing the test area with soap 6 hours 

after application of hexachlorobenzene removed 34% of the applied dose and reduced the cumulative 

amount absorbed after 72 hours by 50% (from 9.71 to 4.90%). 

3.4.2 Distribution 

3.4.2.1 Inhalation Exposure 

Two studies have investigated hexachlorobenzene distribution in people exposed via inhalation. The Aral 

Sea in Uzbekistan is a putative source of airborne exposure to hexachlorobenzene, metals, and other 

pollutants because its water levels are decreasing, revealing sediments to wind dissemination over the 

surrounding area. Ataniyazova et al. (2001) analyzed 18 maternal blood samples, 28 umbilical cord 

blood samples, and 41 milk samples collected from mothers and infants within 200 kilometers of the 

souther border of the Aral Sea in Uzbekistan. The respective mean concentrations of hexachlorobenzene 

in maternal and cord blood were 167 ng/L (range 72–9,920 ng/L) and 70 ng/L (range 25–1,300 ng/L) and 

in 93% of milk samples, with a mean of 28 ng/g fat (range 10–109 ng/L). Sala et al. (2001a) compared 

two populations; for a group of 31 pairs of mothers and infants from Flix, Spain, who had been highly 
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exposed to airborne hexachlorobenzene from a chlorinated solvent factory, the respective geometric 

means of hexachlorobenzene maternal and cord blood levels were 3.98 ng/mL (range 0.50–20.78 ng/mL) 

and 1.40 ng/mL (range 0.30–5.77 ng/mL). For subjects from villages nearby to Flix, the respective 

geometric means of hexachlorobenzene maternal and cord blood levels were 2.51 ng/mL (range 

0.36–7.46 ng/mL) and 0.85 ng/mL (range 0.13–2.45 ng/mL). A statistically significant correlation 

between maternal and umbilical cord blood was detected. 

No studies were located regarding distribution of hexachlorobenzene following inhalation in animals. 

3.4.2.2 Oral Exposure 

Hexachlorobenzene rapidly distributes throughout the body following absorption. A whole-body 

autoradiography experiment in rats showed that hexachlorobenzene was extensively distributed 

throughout the body within 2 hours of receiving a single oral dose of 0.4 mg/kg in peanut oil 

(Ingebrigtsen and Nafstad 1983). In another experiment, radiolabeled hexachlorobenzene was found in 

multiple internal tissues in rats 3 hours after the rats received a single oral dose of about 0.6 mg/kg in 

aqueous methylcellulose (Iatropoulos et al. 1975). 

Although hexachlorobenzene is widely distributed in the body, it is not evenly distributed. Due to its 

lipophilic nature, hexachlorobenzene distributes preferentially to fat, and to a lesser extent, other lipidrich tissues. In the Iatropoulos et al. (1975) study, radiolabel occurred at the highest levels and was most 

persistent in the mesenteric lymph node and adipose tissue. Levels were much lower and declined rapidly 

in the lung, liver, and kidney. The researchers interpreted these findings to indicate that a significant 

fraction of the gastrointestinally absorbed hexachlorobenzene is transported via the lymphatic system to 

the fat, bypassing portal circulation to the liver and systemic circulation to the excretory organs. 

Numerous studies have shown preferential distribution of hexachlorobenzene to adipose tissue. In the rat 

autoradiography study mentioned above, Ingebrigtsen and Nafstad (1983) found peak levels of 

radioactivity in the fat to be roughly 60-fold higher than peak levels in the blood and 30-fold higher than 

peak levels in the brain and liver. Besides the fat, other tissues found to contain relatively high 

concentrations of radioactivity included the skin, bone marrow, Harderian gland, nasal mucosa, preputial 

gland, and intestinal tract. Koss and Koransky (1975) found similar results in rats given single gavage 

doses of 20–180 mg/kg of radiolabeled hexachlorobenzene in oil, with radioactivity in fat two days after 

administration being about 60-fold higher than in blood, 30-fold higher than in liver, and 5-fold higher 
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than in skin. Levels in brain and kidney were intermediate between liver and blood, while levels in 

muscle were lower. Lecavalier et al. (1994) also found approximately 30-fold higher concentrations of 

hexachlorobenzene in the fat than in the liver 14 days after a single gavage dose of 400 or 600 mg/kg of 

hexachlorobenzene in corn oil. Brain and kidney concentrations were slightly (within a factor of 2) lower 

than liver concentrations, while serum levels were an order of magnitude lower than liver concentrations. 

In a repeated dose study, ovariectomized adult female rats treated by gavage in oil with 1, 10, or 

100 mg/kg/day of hexachlorobenzene for 30 days had hexachlorobenzene concentrations in fat that were 

30-fold higher than levels in liver, which were in turn 20-fold higher than levels in serum. Levels in the 

adrenals were roughly 20-fold greater than levels in liver (Foster et al. 1995a). A study in which rats 

were dosed with 50 mg/kg of hexachlorobenzene by gavage in oil every other day for 15 weeks also 

showed hexachlorobenzene concentrations in the fat to be 30–60 times greater than concentrations in the 

liver, brain, kidney, and blood throughout dosing and a subsequent 38-week observation period (Koss et 

al. 1978). 

Preferential distribution to fat and high lipid tissues has also been demonstrated in other animal species. 

Dogs given seven consecutive daily doses of 10 or 100 mg/kg/day of hexachlorobenzene in corn oil by 

capsule had peak hexachlorobenzene concentrations in fat that were 30-fold greater than peak liver levels 

(Sundlof et al. 1982). Other tissues with relatively high levels (5- to 10-fold greater than liver) were the 

skin, adrenal, and thyroid. Levels in the kidney, heart, brain, spleen, pancreas, and muscle were similar to 

levels in the liver. Cynomolgus monkeys treated with 0.1, 1, or 10 mg/kg/day of hexachlorobenzene in 

gelatin capsules for 90 days had hexachlorobenzene concentrations in fat that exceeded liver 

concentrations by 10- to 15-fold (Jarrell et al. 1993). Kidney and brain hexachlorobenzene levels were 

lower than in the liver, while serum and follicular fluid levels were much lower still (Foster et al. 1995b; 

Jarrell et al. 1993). In Rhesus monkeys given oral doses of 8–128 mg/kg/day of hexachlorobenzene by 

gavage in aqueous methyl cellulose for 60 days, the highest concentrations of hexachlorobenzene were 

found in the body fat and bone marrow, with considerably lower concentrations in the adrenals, liver, 

kidney, brain, ovaries, muscle, and serum (Knauf and Hobson 1979). 

The animal studies showed that levels of hexachlorobenzene increased in a dose-dependent manner in all 

tissues, at least at doses up to around 100 mg/kg/day. The study by Foster et al. (1995a) found dosedependent increases in tissue levels of hexachlorobenzene in rats treated with 1, 10, or 100 mg/kg/day, the 

study by Sundlof et al. (1982) found dose-dependent tissue levels in dogs treated with 10 or 

100 mg/kg/day, and the study by Jarrell et al. (1993) found dose-dependent tissue levels in monkeys 

treated with 0.1, 1, or 10 mg/kg/day. In the fat, the increase in hexachlorobenzene concentration was 
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directly proportional to dose at doses up to 100 mg/kg/day in dogs (Sundlof et al. 1982), 10 mg/kg/day in 

rats (Foster et al. 1995a), and 1 mg/kg/day in monkeys (Jarrell et al. 1993). Lecavalier et al. (1994) found 

no difference in tissue levels of hexachlorobenzene following dosing with 400 or 600 mg/kg in rats, 

showing that dose-dependence is lost at high doses. Knauf and Hobson (1979) found no clear 

relationship between tissue levels of hexachlorobenzene and dose in Rhesus monkeys given between 

8 and 128 mg/kg. However, their highly variable results may have been due to very small group sizes 

(one or two monkeys per dose) and, they speculated, variation in the amount of body fat in the monkeys 

used. Although Koss and Koransky (1975) tested multiple dose levels, their results were not presented in 

sufficient detail to assess dose-dependence of tissue levels. 

In humans, data regarding tissue concentrations of hexachlorobenzene are limited to autopsy cases and 

easily sampled tissues and fluids, such as breast milk and blood serum. Two studies were located in 

which both breast adipose/milk and blood serum were collected from the same individuals and analyzed 

for hexachlorobenzene. In a study of 36 Connecticut women between 50 and 80 years of age, hexachlorobenzene was detected in the breast adipose of all 36 women (median concentration=17.7 ng/g fat), 

but was not detected in the serum of any of these women (quantitation limit=0.6 ng/g) (Archibeque-Engle 

et al. 1997). In a study of seven pregnant/lactating New York women (each with a differing interval 

between collection of blood and milk samples), hexachlorobenzene was detected in the serum 

(0.03–0.29 ng/g), but at lower concentrations than in the milk (0.21–0.74 ng/g) (Greizerstein et al. 1999). 

The difference between serum and milk levels can be attributed in part to the differing lipid content of 

these fluids. Lipid-normalized concentrations of hexachlorobenzene were 8–48.4 ng/g lipid in serum and 

11–22.5 ng/g lipid in milk. Two other studies were located that looked at both breast adipose/milk and 

serum levels of hexachlorobenzene in the same populations, although not necessarily in the same 

individuals. In Veracruz, Mexico, a group of 65 volunteer mothers in the hospital for delivery had 

average hexachlorobenzene blood serum concentrations of 1.1 ng/g, and a group of 60 volunteer mothers 

in the hospital for Cesarian delivery (extent of overlap with blood volunteers unknown) had average 

hexachlorobenzene concentrations in milk fat and adipose tissue of 25 and 58 ng/g fat, respectively, 

30 days after delivery (Waliszewski et al. 1999a, 1999b). Similarly, a group of six women from a city in 

northern Germany had a mean serum hexachlorobenzene level of 1.0 ng/g, while breast milk samples 

from seven women from northern Germany in the same year (overlap with blood donors not known) 

showed a mean hexachlorobenzene concentration of about 70 ng/g fat (Petzold et al. 1999). 

Other studies have found hexachlorobenzene in human blood (Rutten et al. 1988; Sala et al. 1999b; 

Schlummer et al. 1998; Siyali 1972), liver (Dewailly et al. 1999; Weistrand and Noren 1998), bone 
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marrow (Bucholski et al. 1996; Scheele et al. 1995), brain (Dewailly et al. 1999), fat (Ansari et al. 1986; 

Dewailly et al. 1999; Lordo et al. 1996; Robinson et al. 1990; Scheele et al. 1995; Siyali 1972; Teufel et 

al. 1990; Weistrand and Noren 1998), and breast milk (Craan and Haines 1998; Czaja et al. 1997; Gladen 

et al. 1999; Gocmen et al. 1989; Lunden and Noren 1998; Newsome and Ryan 1999; Scheele et al. 1995; 

Weisenberg 1986; Wickstrom et al. 1983; and many others). In human plasma, close to half of the hexachlorobenzene present is found in the lipoprotein depleted fraction (containing primarily albumin), while 

the rest is split between the high density (.20%), low density (.20%), and very low density (.10%) 

lipoprotein fractions (Noren et al. 1999). 

Schlummer et al. (1998) found that blood levels of hexachlorobenzene (expressed as ng/g blood lipid) 

varied with age in a group of seven volunteers, ranging from 65 to 82 in four young adult volunteers and 

increasing to 230, 680, and 1,420 in the 53-, 76-, and 81-year-old volunteers, respectively. This result 

shows that hexachlorobenzene accumulates in people as they age. Tissue build up occurs because people 

are continually exposed to hexachlorobenzene in the environment and excretion is slow. In their rat 

experiments, Koss and Koransky (1975) observed an estimated elimination half-time of 8–10 days for 

hexachlorobenzene in the fat and other tissues following administration of a single gavage dose of 

20–180 mg/kg. However, this finding was based on only a 14-day observation period. Koss et al. (1978) 

monitored tissue hexachlorobenzene levels 14 and 38 weeks after a 15-week dosing period (50 mg/kg 

every other day) in rats. Although the researchers could not produce an estimate of biological half-life, 

they found that the rate of elimination decreased over time and speculated that elimination of hexachlorobenzene might continue for years. This issue is discussed in more detail in Section 3.4.4 on Elimination 

and Excretion. 

Human studies have shown that hexachlorobenzene can pass from the mother to the neonate through the 

placenta. In a study of 160 full-term neonates in Germany, hexachlorobenzene was found at an average 

concentration of 2.03 μg/L in 1984/1985 and 0.61 μg/L in 1994/1995 in neonatal blood obtained from an 

unblocked peripheral vein within the first 12 hours of life before the first oral feeding (Lackmann et al. 

1996, 1999). Ando et al. (1985) found hexachlorobenzene in maternal blood, placenta, and neonatal cord 

blood in 36 pregnant Japanese women and their babies. There was a statistically significant correlation 

between the concentration of hexachlorobenzene in the placenta and in cord blood. Waliszewski et al. 

(1999c) reported a statistically significant correlation (R=0.87) between levels of hexachlorobenzene in 

maternal blood serum (mean=1 μg/L, detected in 100% of samples) and umbilical cord serum 

(mean=0.8 μg/L, detected in 98% of samples) in a group of 65 volunteer mothers in Veracruz, Mexico. 

Hexachlorobenzene was also found in the cord blood of all 63 births (geometric mean [GM]=1 μg/L) 
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analyzed in the village of Flix, Spain (Sala et al. 1999a), and in all 656 births (GM=0.04 μg/L) analyzed 

in Quebec, Canada (Rhainds et al. 1999). 

Studies in laboratory animals support the findings in humans that hexachlorobenzene is transferred from 

pregnant mother to the fetus through the placenta. Hexachlorobenzene was found in the fetus and 

placenta of pregnant mice treated with 50 mg/kg/day of hexachlorobenzene by gavage in corn oil on 

days 7–11 of gestation and examined 24 hours after the last dose (Courtney et al. 1976). Follow-up 

studies by these researchers demonstrated that fetal (whole body) and placental hexachlorobenzene 

concentrations increased with dose, with the duration of dosing, and with the day of dosing (dosing later 

in gestation leads to higher levels) in mice, and that results in rats were similar to those in mice (Courtney 

and Andrews 1985; Courtney et al. 1979). Other studies showing transfer of maternal hexachlorobenzene 

to the fetus in rats were reported by Nakashima et al. (1997) and Cripps (1990). Goldey et al. (1990) 

measured maternal and fetal tissue levels of hexachlorobenzene in rats given a total dose of 11 mg/kg 

over a 3-day period 3 weeks prior to breeding. They found that fetal blood and liver concentrations were 

slightly lower than maternal blood concentrations, while fetal brain levels were about half of the maternal 

blood levels. Villeneuve and co-workers measured hexachlorobenzene levels in fetal tissues after 

administering hexachlorobenzene at a series of dose levels. The fetuses of pregnant rats treated with 

5–120 mg/kg/day of hexachlorobenzene by gavage in corn oil on days 6–16 of gestation and sacrificed 

for cesarian section on day 22 of gestation showed dose-related increases in hexachlorobenzene 

concentration in whole fetus, fetal liver, and fetal brain (Villeneuve and Hierlihy 1975). The 

concentration of hexachlorobenzene in fetal liver was about 20–40% of the concentration in maternal 

liver. The concentration in fetal brain was about half that in fetal liver. A similar study in rabbits also 

demonstrated dose-dependent placental transfer of hexachlorobenzene, although in this species, fetal liver 

concentrations of hexachlorobenzene were 2–3-fold higher than maternal liver concentrations (Villeneuve 

et al. 1974a). Fetal brain levels, however, were much lower than fetal liver levels and were also less than 

maternal brain concentrations in rabbits. 

The occurrence of hexachlorobenzene in breast milk of humans is well documented in many populations, 

as noted above. Many of the same animal studies that investigated placental transfer of hexachlorobenzene also studied lactational transfer from mothers to their offspring (Courtney and Andrews 1985; 

Cripps 1990; Goldey et al. 1990; Nakashima et al. 1997). These data are discussed in Section 3.4.4 on 

Elimination and Excretion. 
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3.4.2.3 Dermal Exposure 

No studies were located regarding distribution of hexachlorobenzene following dermal exposure in 

humans. 

Male Fischer 344 rats that received a single dermal dose of approximately 2.5 mg/kg 14C-hexachlorobenzene dissolved in tetrachloroethylene applied to a 4 cm2

 clipped area on the back absorbed only 9.7% 

of the dose; 90.3% of the applied dose remained unabsorbed on the skin after 72 hours. The 

concentration of hexachlorobenzene in the liver and blood increased steadily after dermal application. 

Washing decreased mean hexachlorobenzene concentrations in blood from 263 to 0.128 μg/g and in the 

liver from 0.68 to 0.556 μg/g liver at 72 hours. The authors developed a compartment model based on the 

data collected, for application to a 70-kg worker (Koizumi 1991). 

3.4.3 Metabolism 

3.4.3.1 Inhalation Exposure 

No studies were located regarding metabolism in humans or animals after inhalation exposure to hexachlorobenzene. 

3.4.3.2 Oral Exposure 

Hexachlorobenzene is slowly metabolized to pentachlorophenol by the hepatic cytochrome P-450 system 

(CYP3A1, CYP3A2, CYP3A4 isoforms) (Den Besten et al. 1994; Mehmood et al. 1996; Schielen et al. 

1995a), conjugated with glutathione to yield pentachlorothiophenol, or reductively dechlorinated to form 

pentachlorobenzene (Ingebrigtsen et al. 1986; Koss et al. 1979; Renner 1988). Other metabolites include 

less chlorinated benzenes, chlorophenols, S-conjugated phenols, and benzenes (Den Besten et al. 1994; 

Koss et al. 1986). Pentachlorophenol is subsequently converted to tetrachlorohydroquinone (Mehmood et 

al. 1996; van Ommen et al. 1985). 

Pentachlorobenzene and pentachlorophenol were identified as the major metabolites of 14C-labeled hexachlorobenzene (0.03 mg/kg/day) administered in the diet to Rhesus monkeys for 15 months (Rozman et 

al. 1977). In the urine, approximately 50% of the excreted radiolabel was pentachlorophenol, 25% was 

pentachlorobenzene, and the remaining 25% consisted of unidentified metabolites and unchanged hexa­
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chlorobenzene. Of the excreted radioactivity in the feces, 99% was unchanged hexachlorobenzene, with 

<1% pentachlorobenzene and trace amounts of pentachlorophenol. A subsequent report of a similar study 

in Rhesus monkeys found that fecal excretion consisted of 99% unchanged parent compound, about 1% 

pentachlorobenzene, and traces of pentachlorophenol (Rozman et al. 1978). Urinary metabolites 

consisted of 50–75% pentachlorophenol. The remainder of radioactivity (25–50%) was composed of 

pentachlorobenzene, hexachlorobenzene, and tetrachlorobenzene. Only unchanged parent compound was 

found in the plasma, and the red blood cells contained 95% unchanged parent compound and 5% 

pentachlorophenol. 

A similar metabolic pattern was observed in the rat. Extraction and analysis of fecal radioactivity, which 

accounted for 16% of the administered dose, 7 days after gavage administration of 5 mg/kg of 

14C-labeled hexachlorobenzene in arachis oil to rats did not reveal the presence of metabolites. Urine, 

which contained 0.85% of the administered radiolabel, contained 2,4,5-trichlorophenol, 

pentachlorophenol, and several unidentified chlorinated benzenes (Mehendale et al. 1975). Gas/liquid 

chromatography-mass spectrometry identified 20% of the radioactivity as parental hexachlorobenzene 

together with the metabolites pentachlorothiophenol and pentachlorophenol in isolated perfused rat (male 

Wistar) liver treated with 14C-hexachlorobenzene diluted with unlabeled hexachlorobenzene to yield a 

concentration of 0.1 mg hexachlorobenzene/mL perfusate. Most of the radioactivity was found in the 

perfusate and in the liver; unchanged hexachlorobenzene was responsible for most of the radioactivity. 

Traces of pentachlorothiophenol and pentachlorophenol were identified in the perfusate and the liver, 

respectively (Ingebrigtsen et al. 1981, 1986). A study reported that 98% of biliary radioactivity, which 

constituted 3.6% of the administered dose 48 hours after administration, was in the form of metabolites; 

25% of this radioactivity was glutathione-conjugated pentachlorophenol (Ingebrigtsen et al. 1981). No 

sulfur-containing metabolites of hexachlorobenzene were found in the bile. However, a study of the 

metabolic fate of hexachlorobenzene (particularly as it relates to transformation of hexachlorobenzene 

into any methylthio- and methylsulfonyl-metabolites) in male Wistar rats identified methylthiopentachlorobenzene and 1,4-bis-(methylthio)-2,3,5,6-tetrachlorobenzene as metabolites of hexachlorobenzene 

(Jansson and Bergman 1978). These compounds were excreted to a greater extent than the corresponding 

monosubstituted metabolites. One methylthiotetrachlorobenzene was also found. Pentachlorophenol was 

the only detectable metabolite in blood, liver, urine, or feces of female Wistar rats 38 weeks after the end 

of 15-week gavage exposure to 50 mg/kg/day of hexachlorobenzene (Koss et al. 1978). 

In other rat studies, N-acetyl-S(pentachlorophenyl)cysteine (PCTP-NAC) was the most abundant urinary 

product in female Wistar rats administered dietary hexachlorobenzene in 4% corn oil for 13 weeks 
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resulting in doses of 7.5 or 15 mg/kg/day (Den Besten et al. 1994) or treated twice a week for 35 weeks 

by gavage with 50 mg/kg in olive oil (Rietjens et al. 1995). Other rats in the Den Besten et al. (1994) 

study were similarly administered dietary levels of 0.03 or 0.13% pentachlorobenzene to provide a basis 

for comparison. Pentachlorophenol and tetrachlorohydroquinone were common urinary metabolites of 

both hexachlorobenzene and pentachlorobenzene. Mercaptotetrachlorothioanisole (MTCTA), which was 

excreted as a glucuronide, was also detected in the urine of rats given hexachlorobenzene. Pentachlorophenol, pentachlorothiophenol, 2,3,4,6- and 2,3,5,6-tetrachlorophenol, and pentachlorobenzene were 

identified as metabolites of hexachlorobenzene in another study (Richter et al. 1981). Significant 

sex-related differences were observed, with higher amounts of pentachlorothiophenol observed in the 

livers of female rats. This was accompanied by a slower decrease in hepatic levels of hexachlorobenzene 

in the female rat liver compared to the male liver. Sex differences in the metabolism of hexachlorobenzene in the adult rat have also been observed. After 10 weeks of treatment, the urinary excretion of 

pentachlorophenol, 2,3,5,6-tetrachlorobenzene-1,4-diol, and pentachlorothiophenol was greater in females 

than in males in this study (Rizzardini and Smith 1982). Sex-related differences in biotransformation of 

hexachlorobenzene could account for differences observed in porphyrinogenic activity of hexachlorobenzene in male and female rats (D'Amour and Charbonneau 1992; Richter et al. 1981; Rizzardini and 

Smith 1982). 

Two alternatives have been proposed for the mechanism of conversion of hexachlorobenzene to 

pentachlorophenol: formation of an epoxide or free radical attack on the carbon-chlorine bond. 

Substitution of adjacent carbons with chlorine argues against a mechanism involving an epoxide; either 

cleavage of the bond by free radical attack followed by hydroxylation, or conjugation with glutathione 

seems more plausible. However, if the presence of o- and p-hydroxy derivatives of pentachlorophenol 

could be confirmed, there would be a strong possibility that epoxides are intermediates in the 

dechlorination or hydroxylation of hexachlorobenzene or both (Lui et al. 1976). In vitro studies with 

perfused rat livers demonstrated that 14C-labeled hexachlorobenzene was converted to acidic conjugates 

(45%), while 5% was converted to sulfate or glucuronic acid conjugates (Ingebrigtsen et al. 1986). There 

is evidence that mammalian metabolism of hexachlorobenzene to pentachlorophenol is mediated by the 

hepatic cytochrome P-450 system (CYP3A1, CYP3A2, CYP3A4 isoforms; others) (Den Besten et al. 

1994; Mehmood et al. 1996; Schielen et al. 1995a). 

Several other studies in laboratory animals also identified the following mammalian biotransformation 

products of hexachlorobenzene: pentachlorophenol (Ingebrigtsen et al. 1981; Koss et al. 1976, 1979; Lui 

and Sweeney 1975; van Ommen et al. 1985; Yang et al. 1978); pentachlorothiophenol (D'Amour and 
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3. HEALTH EFFECTS 

Charbonneau 1992; To-Figueras et al. 1992); less chlorinated benzenes, chlorophenols, S-conjugated 

phenols and benzenes (Engst et al. 1976; Koss et al. 1979; Renner 1988; Stewart and Smith 1986); and 

tetrachlorohydroquinone and tetrachlorocatechol (Koss et al. 1976, 1979; Lui et al. 1976; Mehmood et al. 

1996; van Ommen et al. 1985, 1989). The various pathways and metabolites of hexachlorobenzene are 

depicted in Figure 3-3. 

There is some evidence to indicate that metabolism of hexachlorobenzene to pentachlorophenol is 

accelerated in rats fed fish oil, in comparison to lard or soybean oil (Umegaki and Ikegami 1998). Rats 

fed fish oil had significant increases in liver cytochrome P-450, blood pentachlorophenol and 
MAJOR METABOLITES 

Cl OH	 H H

Cl reductive Cl Cl 

dechlorination Cl Cl Cl Cl Cl

Lesser Urine 

mitochondrial + + chlorinated 

benzenes Cl Cl Cl Cl Cl Cl Cl Cl

H (Renner 1998; Cl	 Cl Cl van Ommen et al. 

1989) Pentachlorophenol Pentachlorobenzene Tetrachlorobenzene Hexachlorobenzene (Koss et al. 1986; (Mehendale et al. 1975; (Jansson and Bergman
 

Mehendale et al. 1975; Renner 1988; 1978;
 

Richter et al. 1981; Richter et al. 1981; Renner 1988)
 

Rozman et al. 1977a) Rozman et al. 1977a)
 

MINOR METABOLITES 

Cl	 OH OH OH OH 

Cl Cl Cl Cl Cl 	 Cl Cl Cl Cl

monooxygenase epoxyhydrates + + + Urine 

Cl Cl Cl H Cl Cl Cl H Cl 

Cl	 Cl Cl H OH 

Hexachlorobenzene 2,4,5-Trichlorophenol Tetrachlorohydroquinone 2,3,4,6-Tetrachlorophenol 2,3,5,6-Tetrachlorophenol 

(Mehendale et al. 1975)	 (Lui et al. 1989; (Richter et al. 1981) (Richter et al. 1981) 

To-Figueras et al. 1991 

van Ommen et al. 1989) 

GLUTATHIONE CONJUGATION: FORMATION OF SULFUR DERIVATIVES
 

(Adapted from Ingebrigtsen et al. 1981; Jansson and Bergman 1978; Renner 1988; To-Figueras et al. 1992)
 

Cl GSH	 NAC 

Cl Cl Cl Cl Cl Cl 

monooxygenase -Gly acetylation 

Urine 

glutathione-s-transferase 

Cl Cl Cl Cl Cl Cl 

Cl Cl	 Cl 

Hexachlorobenzene Reduced glutathione conjugate N-acetyl cysteine derivatives 

hydrolysis 

SOH SCH3

Cl Cl Cl Cl

Urine derivatization 

Cl Cl Cl Cl

Cl Cl

Pentachlorothiophenol (PCThP)	 Pentachlorothioanisole (PCTA) 

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Figure 3-3. Metabolism and Urinary Metabolites of Hexachlorobenzene 
HEXACHLOROBENZENE 140 

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pentachlorophenol:hexachlorobenzene ratio, and urinary excretion of pentachlorophenol, while levels of 

hexachlorobenzene in the feces were unchanged (indicating no difference in absorption between groups). 

Limited human data are consistent with results from animal studies. Portions (0.5 g) of abdominal 

subcutaneous adipose tissue obtained as histological samples during surgery of patients and urine from 

these patients in Germany were extracted with benzene for hexachlorobenzene and its metabolites (Koss 

et al. 1986). Hexachlorobenzene was detected in the adipose tissue of the patients, while only 

pentachlorophenol was detected in the urine. A correlation was found between levels of hexachlorobenzene found in adipose tissue and urinary pentachlorophenol. However, it is possible that the urinary 

pentachlorophenol originated from other chlorinated hydrocarbons such as pentachlorobenzene, 

gamma-hexachlorocyclohexane, or pentachloronitrobenzene. Human cytochrome P-450 3A4 expressed 

in the yeast Saccharomyces cerevisiae metabolized hexachlorobenzene to pentachlorophenol, which was 

further transformed to tetrachlorohydroquinone, in both in vitro and in vivo experiments (Mehmood et al. 

1996). 

3.4.3.3 Dermal Exposure 

No studies were located regarding metabolism in humans or animals after dermal exposure to hexachlorobenzene. 

3.4.4 Elimination and Excretion 

3.4.4.1 Inhalation Exposure 

To-Figueras et al. (2000) observed a high correlation between fecal and blood levels of hexachlorobenzene in a group of 25 men and 28 women from Flix, Spain. This population was highly exposed to 

airborne hexachlorobenzene from a nearby chlorinated solvent factory. The geometric mean of hexachlorobenzene in blood was 163 μg/5.4 L. Estimated fecal excretion of unchanged hexachlorobenzene 

was 10.4 μg/day, 4–6.4% of the estimated total blood level. No unchanged hexachlorobenzene was 

detected in urine; urinary excretion of metabolites was 5.1 μg/day, 1.8–3.1% of the estimated total blood 

level. 

No studies were located regarding excretion of hexachlorobenzene in animals following inhalation 

exposure. 
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3.4.4.2 Oral Exposure 

Elimination of absorbed hexachlorobenzene is slow and occurs primarily via the feces, with smaller 

amounts being excreted in the urine. Hexachlorobenzene eliminated in the feces is predominantly 

unchanged parent compound, although small amounts of various metabolites have also been found. 

Conversely, hexachlorobenzene in the urine is almost all in the form of metabolites. Following a single 

intraperitoneal injection of radiolabeled hexachlorobenzene (4 mg/kg) in rats, approximately 34% of the 

administered radioactivity was recovered in the feces over the following 14 days, 80% of which was 

unchanged parent compound (Koss and Koransky 1975). By contrast, only 5% of the administered 

radioactivity was recovered in the urine, and only 4% of that was unchanged parent compound. A 14-day 

recovery period was used due to the slow elimination of hexachlorobenzene from the body. Yang et al. 

(1978) conducted an experiment in which rats were injected intravenously with radiolabeled hexachlorobenzene and excretia collected over the following 48 hours. Only 1% of the administered dose appeared 

in the feces, and 0.2% in the urine, within 48 hours of dosing. These researchers also conducted a longterm experiment in monkeys (Yang et al. 1978). One year after intravenous injection, only 28.2% of the 

administered dose had been excreted in the feces (90% unchanged hexachlorobenzene) and 1.6% in the 

urine (all metabolites). The researchers attributed slow elimination of hexachlorobenzene to long-term 

storage of the chemical in the fat. Neither Koss and Koransky (1975) nor Yang et al. (1978) detected 

radioactivity in the expired air from treated animals. 

Both biliary and intestinal excretion contribute to fecal excretion of hexachlorobenzene. Hexachlorobenzene and/or its metabolites were found in the bile after intravenous injection of hexachlorobenzene in 

monkeys, suggesting the possibility of biliary excretion (Yang et al. 1978). Bile duct cannulated rats 

given hexachlorobenzene by gavage excreted 3.6% of the administered dose in the bile within 48 hours 

(Ingebrigtsen et al. 1981). Although one report suggested that biliary excretion was more important than 

intestinal excretion (Sundlof et al. 1982), other studies have shown that biliary excretion is a minor 

contributor to fecal excretion. Rozman et al. (1981) estimated biliary excretion to account for about 10% 

of fecal excretion in rats and monkeys treated with oral hexachlorobenzene. Intestinal excretion was 

responsible for the bulk of the fecal excretion in this study. Feeding with aliphatic hydrocarbons (mineral 

oil, hexadecane) enhanced fecal excretion in both rats and monkeys, with a corresponding decrease in 

blood and adipose hexachlorobenzene concentrations, primarily due to increased elimination of hexachlorobenzene in the large intestine (Rozman et al. 1981). In contrast to aliphatic hydrocarbons, 

cholestyramine, which interferes with enterohepatic recycling, had no effect on fecal excretion of hexachlorobenzene (confirming the minor role of biliary excretion for this chemical), and sesame oil produced 
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only a very small increase in fecal hexachlorobenzene excretion (possibly by increasing gastrointestinal 

absorption). Richter and Schafer (1981) showed that addition of hydrocarbons (light liquid paraffin and 

squalane) to the perfusion medium enhanced elimination of unchanged hexachlorobenzene in perfused 

intestine into the lumen of the jejunum, ileum, and colon. The researches hypothesized that the 

hydrocarbons, which are not significantly absorbed, act as a lipophilic compartment in the gut lumen, 

shifting the equilibrium between gut wall and lumen in favor of the lumen for hydrophilic substances 

such as hexachlorobenzene. This is consistent with the fat-flush hypothesis of gastrointestinal absorption 

proposed by Schlummer et al. (1998). After fat-flush enhanced lipid absorption in the duodenum and 

jejunum is complete, the diffusion gradient is reversed in subsequent portions of the intestines. 

In lactating mothers, breast milk is also an important route of excretion for hexachlorobenzene. Hexachlorobenzene has been detected in human breast milk in many studies (Craan and Haines 1998; Czaja et 

al. 1997; Gladen et al. 1999; Gocmen et al. 1989; Lunden and Noren 1998; Newsome and Ryan 1999; 

Scheele et al. 1995; Weisenberg 1986; Wickstrom et al. 1983; and many others). Several animal studies 

have quantified lactational transfer of hexachlorobenzene from mothers to their offspring (Courtney and 

Andrews 1985; Cripps 1990; Goldey et al. 1990; Nakashima et al. 1997). These studies confirm the 

importance of breast milk as a route of elimination in the mother and as a source of exposure in neonates. 

Studies that monitored elimination of hexachlorobenzene for an extended period of time noted that the 

rate of elimination decreases over time (Koss et al. 1978, 1983; Sundlof et al. 1982; Yang et al. 1978). In 

rats treated with hexachlorobenzene every other day for 6 weeks, the elimination half-time was estimated 

as a relatively rapid 8 days soon after exposure stopped, a much slower 10 weeks 3 months later, and a 

very slow 12 months after 1.5 years, suggesting that elimination of hexachlorobenzene could continue for 

years (Koss et al. 1978, 1983). Yang et al. (1978) and Sundlof et al. (1982) both applied 3-compartment 

pharmacokinetic models to their data on dogs and monkeys, respectively. Sundlof et al. (1982) obtained 

elimination half-time estimates ranging from 6 weeks to 3 years in the three dogs modeled. Yang et al. 

(1978) calculated elimination rate constants corresponding to half-times of 91–114 days in two monkeys, 

but also performed additional modeling exercises that suggested elimination half times as long as 

250 years. Freeman et al. (1989) used a physiologically based pharmacokinetic (PBPK) model of hexachlorobenzene in the rat to show that approximately 75% of the decline in fat concentrations over time is 

due to growth (i.e., dilution), with only 25% due to excretion. Thus, the growth of animals during 

experiments may affect the apparent half-life of hexachlorobenzene. A PBPK model developed by Yesair 

et al. (1986) predicted a half-life of 215 days for hexachlorobenzene in a growing human female exposed 

to doses of 0.5–32 mg/kg/day for 15 weeks at 15 years of age. 
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3.4.4.3 Dermal Exposure 

No studies were located regarding excretion in humans or animals after dermal exposure to hexachlorobenzene. 

3.4.4.4 Other Routes of Exposure 

No studies were located regarding excretion in humans after other routes of exposure to hexachlorobenzene. 

The major portion of injected hexachlorobenzene is eliminated unchanged in feces, while a smaller 

fraction, composed of metabolites, is eliminated in urine. Yang et al. (1978) administered a single 

intravenous dose of hexachlorobenzene to monkeys; 1 year after intravenous injection, only 28.2% of the 

administered dose had been excreted in the feces, 90% unchanged, and in urine, 1.6% of the total dose 

was excreted as metabolites (no unchanged compound detected). The researchers attributed slow 

elimination of hexachlorobenzene to long-term storage of the chemical in the fat. Unchanged hexachlorobenzene and metabolites were detected in bile. Yang et al. (1978) administered a single intravenous dose 

of hexachlorobenzene to rats; within 48 hours, only 1% appeared in the feces, and 0.2% in the urine. 

Following a single intraperitoneal injection of hexachlorobenzene, rats excreted 34% of the total dose in 

feces over 14 days, 80% of which was unchanged parent compound; only 5% was excreted in urine over 

14 days and only 4% was unchanged parent compound (Koss and Koransky 1975). Neither Koss and 

Koransky (1975) nor Yang et al. (1978) detected hexachlorobenzene elimination in air expired by treated 

animals. Yang et al. (1978) calculated an initial half-life of 91–114 days, and subsequent half-lives as 

long as 250 years. 

3.4.5 Physiologically Based Pharmacokinetic (PBPK)/Pharmacodynamic (PD) Models 

Physiologically based pharmacokinetic (PBPK) models use mathematical descriptions of the uptake and 

disposition of chemical substances to quantitatively describe the relationships among critical biological 

processes (Krishnan et al. 1994). PBPK models are also called biologically based tissue dosimetry 

models. PBPK models are increasingly used in risk assessments, primarily to predict the concentration of 

potentially toxic moieties of a chemical that will be delivered to any given target tissue following various 

combinations of route, dose level, and test species (Clewell and Andersen 1985). Physiologically based 
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pharmacodynamic (PBPD) models use mathematical descriptions of the dose-response function to 

quantitatively describe the relationship between target tissue dose and toxic end points. 

PBPK/PD models refine our understanding of complex quantitative dose behaviors by helping to 

delineate and characterize the relationships between: (1) the external/exposure concentration and target 

tissue dose of the toxic moiety, and (2) the target tissue dose and observed responses (Andersen and 

Krishnan 1994; Andersen et al. 1987). These models are biologically and mechanistically based and can 

be used to extrapolate the pharmacokinetic behavior of chemical substances from high to low dose, from 

route to route, between species, and between subpopulations within a species. The biological basis of 

PBPK models results in more meaningful extrapolations than those generated with the more conventional 

use of uncertainty factors. 

The PBPK model for a chemical substance is developed in four interconnected steps: (1) model 

representation, (2) model parametrization, (3) model simulation, and (4) model validation (Krishnan and 

Andersen 1994). In the early 1990s, validated PBPK models were developed for a number of 

toxicologically important chemical substances, both volatile and nonvolatile (Krishnan and Andersen 

1994; Leung 1993). PBPK models for a particular substance require estimates of the chemical substancespecific physicochemical parameters, and species-specific physiological and biological parameters. The 

numerical estimates of these model parameters are incorporated within a set of differential and algebraic 

equations that describe the pharmacokinetic processes. Solving these differential and algebraic equations 

provides the predictions of tissue dose. Computers then provide process simulations based on these 

solutions. 

The structure and mathematical expressions used in PBPK models significantly simplify the true 

complexities of biological systems. If the uptake and disposition of the chemical substance(s) is 

adequately described, however, this simplification is desirable because data are often unavailable for 

many biological processes. A simplified scheme reduces the magnitude of cumulative uncertainty. The 

adequacy of the model is, therefore, of great importance, and model validation is essential to the use of 

PBPK models in risk assessment. 

PBPK models improve the pharmacokinetic extrapolations used in risk assessments that identify the 

maximal (i.e., the safe) levels for human exposure to chemical substances (Andersen and Krishnan 1994). 

PBPK models provide a scientifically sound means to predict the target tissue dose of chemicals in 

humans who are exposed to environmental levels (for example, levels that might occur at hazardous waste 
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sites) based on the results of studies where doses were higher or were administered in different species. 

Figure 3-4 shows a conceptualized representation of a PBPK model. 

3.4.5.1 Summary of PBPK Models 

PBPK models for hexachlorobenzene have been developed by Yesair et al. (1986) and Freeman et al. 

(1989). The Yesair et al. (1986) model describes the absorption, distribution, and elimination of ingested 

hexachlorobenzene in growing rats and humans. The Freeman et al. (1989) model describes distribution 

and excretion of intravenously injected hexachlorobenzene in growing rats. 

3.4.5.2 Hexachlorobenzene PBPK Model Comparison 

The Yesair et al. (1986) and Freeman et al. (1989) models are similar attempts to describe distribution and 

clearance of hexachlorobenzene. Both models included numerous tissue compartments and allowed for 

growth over time. The Yesair et al. (1986) model went further than the Freeman et al. (1989) model by 

including oral exposure, fetal and breast milk compartments, and metabolism and tissue sequestration of 

hexachlorobenzene, and by modeling humans as well as rats. Both rat models were validated using 

experimental data. 
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3. HEALTH EFFECTS 

Figure 3-4. Conceptual Representation of a Physiologically
 

Based Pharmacokinetic (PBPK) Model for a 

Hypothetical Chemical Substance
 

Source: adapted from Krishnan et al. 1994 

Note: This is a conceptual representation of a physiologically based pharmacokinetic (PBPK) model for a hypothetical 

chemical substance. The chemical substance is shown to be absorbed via the skin, by inhalation, or by ingestion, 

metabolized in the liver, and excreted in the urine or by exhalation. 
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3. HEALTH EFFECTS 

3.4.5.3 Discussion of Models 

The Yesair Model. 

Risk assessment. The Yesair model was not used for risk assessment. 

Description of the model. The Yesair model was initially developed to simulate oral exposure to 

hexachlorobenzene in growing male and female rats, and was then expanded to include pregnancy and 

offspring in the female model. A human model was produced by using the same model structure with 

human physiological parameter values. The model includes compartments for intestinal lumen, systemic 

circulation, feces, liver, metabolites, kidney, urine, brain, richly perfused tissues, poorly perfused tissues, 

breast milk, fetus, and lactating offspring. Parameters used in the model included body and organ weights 

and growth rates, blood-flow rates, empirical clearance factors, reaction-rate constants, distribution ratios, 

and capacity limits. The forms of the growth characteristics and the parameter values were obtained from 

the literature, experimental data from Kuiper-Goodman et al. (1977), and empirical considerations. Both 

free and sequestered forms of hexachlorobenzene were estimated in each compartment, and only freely 

available material was allowed to leave the compartment (except for the systemic circulation and breast 

milk compartments). When the rat models were adapted for humans, the appropriate human 

physiological values were substituted for the rat values. 

Validation of the model. The rat model was compared to data from Courtney et al. (1979), 

Iatropoulos et al. (1975), Koss and Koransky (1975), Koss et al. (1978), and Kuiper-Goodman et al. 

(1977). In general, the model approximated the observed results reasonably well in all tissues. The 

human model was not validated. 

Target tissues. Using the rat model, correlations were obtained between predicted yield of 

metabolites and experimentally observed liver toxicity (porphyria and increased liver weight reported by 

Koss et al. 1978 and Kuiper-Goodman et al. 1977), between estimated concentrations of hexachlorobenzene in the brain and observed central nervous system effects (tremors and other signs reported by 

Koss et al. 1978 and Kuiper-Goodman et al. 1977), and between estimated extent of lactational transfer of 

hexachlorobenzene and observed offspring mortality (data from Kitchen et al. 1982). The model predicts 

little transfer of hexachlorobenzene to the fetus during gestation and extensive mobilization of hexachlorobenzene to the offspring during lactation. The data are consistent with this model. A half-life of 

215 days was predicted for hexachlorobenzene in a growing human female exposed to doses of 
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0.5–32 mg/kg/day for 15 weeks at 15 years of age, suggesting that approximately 4 years (7 half-lives) 

are required to establish equilibrium between intake and excretion. Further simulations showed that 

doubling or halving the administered dose resulted in doubling or halving, respectively, of the tissue 

concentrations after 3–5 years. 

Species extrapolation. Species extrapolation was not attempted in this model. 

Interroute extrapolation. Interroute extrapolation was not attempted in this model. 

The Freeman Model. 

Risk assessment. The Freeman model was not used for risk assessment. 

Description of the model. The Freeman model was developed to simulate intravenous injection of 

hexachlorobenzene in growing rats. The model includes compartments for plasma, gastrointestinal tract, 

colon, feces, liver, lung, kidney, urine, brain, heart, spleen, skin, muscle, and fat. Parameters used in the 

model include organ weights and blood flow fractions obtained from the literature and tissue:serum 

partition coefficients derived from experimental studies by Scheufler and Rozman (1984a, 1984b). The 

model was designed to accommodate differential growth of tissue/organ weights as a function of total 

body weight. Metabolism was assumed to be zero based on experimental data (attributed to Rozman and 

colleagues) suggesting little metabolism of hexachlorobenzene in the rat. 

Validation of the model. The model predictions were compared to data from Scheufler and Rozman 

(1984a, 1984b). In general, the model approximated the observed results reasonably well in the 

compartments examined: blood, liver, fat, urine, and feces. 

Target tissues. Levels in liver and fat were well-predicted by this model. An interesting prediction is 

that approximately 75% of the decline in fat concentrations over time is due to growth (i.e., dilution), with 

only 25% due to excretion. Thus, the growth of animals during experiments may affect the apparent halflife of hexachlorobenzene. 
HEXACHLOROBENZENE 149 

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Species extrapolation. Species extrapolation was not attempted in this model. 

Interroute extrapolation. Interroute extrapolation was not attempted in this model. 

3.5 MECHANISMS OF ACTION 

3.5.1 Pharmacokinetic Mechanisms 

Absorption. Schlummer et al. (1998) found that absorption following oral exposure to hexachlorobenzene decreased with increasing hexachlorobenzene blood levels in volunteers. Data was fit to a linear 

regression. The calculated regression equation was: 

percent absorption hexachlorobenzene = 0.8538–0.0021 x (ng hexachlorobenzene/g blood lipid) 

This equation indicates that 85.4% of ingested hexachlorobenzene will be absorbed when the blood 

contains no hexachlorobenzene, and that this percentage will be reduced by 0.2% for each ng of hexachlorobenzene per g lipid in blood. This regression equation provided an excellent fit to the data 

(R2

=0.98), leading to the conclusion that concentration of hexachlorobenzene in the blood is a major 

factor influencing absorption of hexachlorobenzene from the gut. Although absorption of lipophilic 

compounds such as hexachlorobenzene is thought to occur by simple passive diffusion, analysis of the 

entire data set in this study, which included 17 polychlorinated dibenzo-p-dioxins/furan congeners and 

10 PCBs in addition to hexachlorobenzene, showed that absorption of these compounds could not be 

explained by simple gradients between lipid-based food and blood concentrations or between gut contents 

and gut tissue. The researchers offered a “fat-flush” hypothesis of absorption for hexachlorobenzene and 

related compounds whereby lipid absorbed into gut tissue would temporarily increase lipid content of the 

tissue, diluting contaminant concentration in the gut tissue (on a lipid basis) and increasing the diffusion 

gradient for the contaminant from the gut contents, before being transported from the gut tissue to the 

lymph and blood. 

Animal studies have demonstrated that hexachlorobenzene is absorbed more rapidly from oil than from 

aqueous suspension (Albro and Thomas 1974; Iatropoulos et al. 1975; Ingebrigtsen and Nafstad 1983; 

Ingebrigtsen et al. 1981; Koss and Koransky1975; Sundlof et al. 1982; Zabick and Schemmel 1980; and 

others). Nakashima et al. (1999) observed that lactational transfer of hexachlorobenzene was increased 

by feeding nursing rats high fat diets. Zabick and Schemmel (1980) demonstrated that a high fat diet also 
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3. HEALTH EFFECTS 

enhances absorption of hexachlorobenzene, in comparison to a high carbohydrate diet. Groups of six 

female Osbourne-Mendel rats were fed either a high fat diet or one of two high carbohydrate diets 

supplemented with 32 mg/kg/day of hexachlorobenzene for 6–18 days. The high fat diet resulted in 

higher carcass fat content and greatly increased concentrations of hexachlorobenzene in perirenal fat, 

liver, and gastrocnemius. At the same time, concentrations of hexachlorobenzene in the feces were much 

lower with the high fat diet, suggesting that the high fat diet facilitated absorption of hexachlorobenzene 

from the gut, thereby leading to increased tissue levels. These data showing enhanced absorption of 

hexachlorobenzene when administered in oil or a high fat diet are consistent with the “fat-flush” 

hypothesis for hexachlorobenzene absorption proposed by Schlummer et al. (1998). 

Iatropoulos et al. (1975) measured absorption in Sprague-Dawley rats following oral gavage of hexachlorobenzene in aqueous methylcellulose suspension, and observed that hexachlorobenzene was 

absorbed through the walls of the stomach, duodenum, jejunum, and ileum. Based on their data, the 

authors concluded that the majority of ingested hexachlorobenzene was absorbed from the small intestine 

by the lymphatic system and deposited in the fat, bypassing portal circulation (to the liver), systemic 

circulation, and the excretory organs. 

Distribution. Hexachlorobenzene has been detected in human and animal blood, and circulation is the 

primary mechanism of inter-tissue distribution. Due to its lipophilic nature, hexachlorobenzene 

distributes preferentially to fat, and to a lesser extent, other lipid-rich tissues (Foster et al. 1995a, 1995b; 

Iatropoulos et al. 1975; Ingebrigtsen and Nafstad 1983; Jarrell et al. 1993; Knauf and Hobson 1979; Koss 

and Koransky 1975; Koss et al. 1978; Lecavalier et al. 1994; Sundlof et al. 1982). Although distribution 

studies have produced somewhat conflicting results, a rough approximation of hexachlorobenzene 

distribution is: fat > lymph nodes, adrenal and thyroid glands > skin > bone marrow, Harderian gland, 

nasal mucosa, preputial gland, and intestinal tract > liver, lung > kidney, brain, blood, heart, muscle, 

spleen, pancreas, ovary. 

Human and animal studies suggest that breast milk is enriched with hexachlorobenzene, relative to blood, 

and that blood levels actually drop in lactating mothers (Greizerstein et al. 1999; Petzold et al. 1999; 

Nakashima et al. 1997, 1999; Nakashima and Ikegari 2000; Waliszewski et al. 1999a, 1999b). This is 

probably due to the lipophilicity of hexachlorobenzene. 
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Metabolism. Hexachlorobenzene is slowly metabolized by hepatic cytochrome P-450 system 

(CYP3A1, CYP3A2, CYP3A4 isoforms) (Den Besten et al. 1994; Mehmood et al. 1996; Schielen et al. 

1995a); by conjugation with glutathione, glucuonide, and sulfate; and by reductive dechlorination 

(Ingebrigtsen et al. 1986; Koss et al. 1979; Renner 1988). 

Two alternatives have been proposed for the mechanism of conversion of hexachlorobenzene to PCP: 

formation of an epoxide or free radical attack on the carbon-chlorine bond. Substitution of adjacent 

carbons with chlorine argues against a mechanism involving an epoxide; either cleavage of the bond by 

free radical attack followed by hydroxylation, or conjugation with glutathione seems more plausible. 

However, if the presence of o- and p-hydroxy derivatives of PCP could be confirmed, there would be a 

strong possibility that epoxides are intermediates in the dechlorination or hydroxylation of hexachlorobenzene or both (Lui et al. 1976). 

Excretion. Following metabolism to more polar metabolites, hexachlorobenzene is excreted in urine 

(Ingebrigtsen et al. 1981, 1986; Koss and Koransky 1975; Koss et al. 1986; Lui and Sweeney 1975; 

Rozman et al. 1977a; Scheufler and Rozman 1984a, 1984b; To-Figueras et al. 1992; van Ommen et al. 

1985; Yang et al. 1978). 

Hexachlorobenzene eliminated in the feces is predominantly unchanged parent compound, although small 

amounts of various metabolites have also been found (Koss and Koransky 1975; Yang et al. 1978). Fecal 

elimination is primarily the product of fecal excretion, although biliary excretion (from the liver) is also 

important (Ingebrigtsen et al. 1981; Richter and Schafer 1981; Rozman et al. 1981; Yang et al. 1978). 

For nursing mothers, excretion of unchanged hexachlorobenzene into milk may represent a significant, 

and even the primary, route of excretion (Courtney and Andrews 1985; Craan and Haines 1998; Cripps 

1990; Czaja et al. 1997; Gladen et al. 1999; Gocmen et al. 1989; Goldey et al. 1990; Lunden and Noren 

1998; Nakashima et al. 1997; Newsome and Ryan 1999; Scheele et al. 1995; Weisenberg 1986; 

Wickstrom et al. 1983; and many others). 
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3.5.2 Mechanisms of Toxicity 

Hexachlorobenzene induces porphyria characterized by increased d-ALA synthase (the enzyme that 

controls the rate of porphyrin production) activity and decreased uroporphyrinogen decarboxylase (the 

enzyme that converts uroporphyrinogen III to coproporphyrinogen III) activity (Dowdele et al. 1967; 

Rajamanickam et al. 1972; Smith and de Matteis 1990; Strand et al. 1971). Uroporphyrinogen III is the 

first cyclic tetrapyrrole in the pathway of heme biosynthesis. This is the reduced colorless precursor of 

uroporphyrin III (hexahydro-uroporphyrin) and will give rise to the corresponding porphyrin on 

reoxidation. Uroporphyrinogen decarboxylase (a cytosolic enzyme) converts uroporphyrinogen III to 

coproporphyrinogen III by the stepwise decarboxylation of the four acetic acid side chains to leave 

methyl residues, but the corresponding porphyrin (uroporphyrin III) cannot be decarboxylated and will 

not be metabolized further. Thus, the accumulation of uroporphyrins in the liver may be due to a 

deficiency of the decarboxylation of uroporphyrinogen III catalyzed by uroporphyrinogen decarboxylase. 

This hypothesis led to the proposal that in certain porphyria where uroporphyrin accumulates 

(uroporphyrias), the mechanism responsible may be an accelerated oxidation of uroporphyrinogen, 

causing an "oxidative escape" of this intermediate from the pathway of heme biosynthesis (Meola and 

Lim 1993; Smith and de Matteis 1990). A marked inhibition of uroporphyrinogen decarboxylase has 

been widely reported to occur prior to the manifestation of the typical porphyrinogenic effects of hexachlorobenzene (Blekkenherst et al. 1976; Elder et al. 1976; Smith and Francis 1987; Smith et al. 1986a). 

A study conducted with Osteogenic Disorder Shionogi (ODS) rats provided evidence that chemicallyinduced porphyria seems to be mediated by inhibition of CYP1A2-catalyzed uroporphyrinogen oxidation 

(Sinclair et al. 1995). 

In vitro studies with hexachlorobenzene have demonstrated that this chemical does not exert a direct 

action on uroporphyrinogen decarboxylase (Rios de Molina et al. 1980). The major hexachlorobenzene 

metabolites (TCHQ, PCP, pentachlorothiophenol, pentachlorothioanisole) had no influence on the 

porphyrin pathway as indicated by alteration in total hepatic porphyrin levels and urinary levels of 

d-ALA, porphobilinogen, or uroporphyrins (Goldstein et al. 1977, 1978; Kimbrough and Linder 1978; 

Lui et al. 1976; Smith and Francis 1987; Wainstok de Calmanovici and San Martin de Viale 1980). 

However, PCP and TCHQ appear to be capable of altering porphyrin metabolism in in vitro systems 

containing d-ALA (Goerz et al. 1978; Goldstein et al. 1977). Furthermore, co-administration of PCP and 

TCHQ with hexachlorobenzene increased the severity of the resultant porphyria (Debets et al. 1980a), 

indicating a probable role for these metabolites in porphyria induction. It has been suggested that changes 

in K+

 permeability mediated by lipid peroxidation and mitochondrial dysfunction may be contributing 
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factors in hexachlorobenzene-induced hepatotoxicity. This was based on the results of a study in rats in 

which mitochondrial lipid peroxidation was found to have increased proportionally with a 100-fold 

increase in hepatic porphyrin content (Feldman and Bacon 1989; Masini et el. 1988). Porphyrin uptake in 

the mitochondria of iron-supplemented rats was inhibited by PCP (hexachlorobenzene metabolite), 

indicating that peroxidative reactions in the mitochondrial membranes may be responsible for changes in 

membrane permeability (Masini et al. 1988). 

It is unlikely, however, that the hexachlorobenzene metabolites, pentachlorobenzene and PCP, are by 

themselves porphyrinogenic agents, and P-450 induction may not correlate with porphyria development. 

In one study, rats fed diets containing hexachlorobenzene or its metabolites (pentachlorobenzene and 

PCP) exhibited increases in hepatic cytochrome P-450, but the metabolites had no effect on urinary 

porphyrin excretion, while hexachlorobenzene produced a high level of urinary porphyrins (Vos et al. 

1988). In other studies, the major hexachlorobenzene metabolites (TCHQ, PCP, pentachlorothiophenol, 

pentachlorothioanisole) had no influence on the porphyrin pathway as indicated by alteration in total 

hepatic porphyrin levels and urinary levels of d-ALA, porphobilinogen, or uroporphyrins (Goldstein et al. 

1977, 1978; Kimbrough and Linder 1978; Smith and Francis 1987; Wainstok de Calmanovici and San 

Martin de Viale 1980). However, PCP and TCHQ appear to be capable of altering porphyrin metabolism 

in in vitro systems containing d-ALA (Goerz et al. 1978; Goldstein et al. 1977). Furthermore, coadministration of PCP and TCHQ with hexachlorobenzene increased the severity of hexachlorobenzeneinduced porphyria (Debets et al. 1980a). 

In a study with rats, it was proposed that the involvement of the histidine residue of the enzyme in 

substrate (hexachlorobenzene) binding may be the mechanism by which hexachlorobenzene exerts its 

porphyrinogenic action in vivo (Billi de Catabbi et al. 1991). Another study in rats and mice concluded 

that hexachlorobenzene induces chronic porphyria by modifying sulfhydryl groups in porphyrinogen 

decarboxylase, the action restricted to the catalytic or substrate-binding sites (Elder and Urquhart 1986). 

Iron (as iron dextran) alone has been shown to induce porphyria; therefore, iron may have a role in the 

pathogenesis of hexachlorobenzene-induced porphyria (Siersema et al. 1991; Smith and Francis 1983; 

Smith et al. 1986a) and hepatocarcinogenesis (Smith 1989), although co-administration of carbonyl iron 

did not have a significant effect on elevated hepatic and mitochondrial fraction porphyrin contents in rats 

following hexachlorobenzene treatment (Masini et al. 1988), but in other animal studies, exposure of mice 

to hexachlorobenzene and iron produced a dramatic increase (nearly 1,000-fold) in hepatic uroporphyrin 

levels (Vincent et al. 1989). The small number of animals used in this study limits the reliability of the 
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conclusions of this study. Another investigator concluded that liver mitochondrial porphyrin uptake may 

involve the K+

 transmembrane gradient and further suggested that peroxidative reactions in the 

mitochondrial membranes may be responsible for changes in membrane permeability that affects K+ 

permeability (Masini et al. 1988). Liver malonaldehyde levels increased while glucose-6-phosphate 

activity decreased in rats administered intraperitoneal injections of 50 mg of hexachlorobenzene per day 

(total dose=300 mg) during a 42-day test period followed by treatment with intraperitoneal doses of iron 

(as ferrihydroxide-dextran complex), suggesting a close relationship between accumulation of porphyrins, 

iron overload, and free radical formation or lipid peroxidation (Visser et al. 1989). 

Experiments with rats and iron-loaded mice indicate that there may also be an association between the 

induction of uroporphyria and the development of liver tumors after the administration of polyhalogenated aromatic chemicals (Smith and De Matteis 1990). Hyperplastic nodules were observed in the 

liver lobes of 80% of female Fischer 344 rats pretreated with Imferon (iron-dextran) and then given 

dietary hexachlorobenzene in arachis oil for 65 weeks resulting in a dose of 5 or 10 mg/kg/day. There 

was a high incidence of fibrin throughout the liver with sinusoidal telangiectasis, and nodular peliosis 

hepatitis and hepatocellular necrosis. The study proposed a nongenotoxic mechanism for tumor induction 

by hexachlorobenzene, concluding that the formation of hepatomas and hemangiomas with elements of 

peliosis could be explained by the compensatory hyperplastic responses to hepatocellular injury or 

necrosis and by the simultaneous loss of hepatocellular cords. The study further concluded that the 

accumulation of iron in the liver would strongly potentiate the development of hepatic tumors (Carthew 

and Smith 1994). 

Iron overload also greatly sensitized mice to the development of liver tumors. Mice given oral hexachlorobenzene doses preceded by subcutaneous administration of iron developed iron-excluded 

hyperplastic nodules (all treated animals) and hepatocellular carcinoma (most animals). Based on the 

results of this investigation, an alternate mechanism has been suggested for the hepatic toxicity of hexachlorobenzene that may involve the uncoupling of an induced cytochrome P-450 system releasing active 

oxygen species. Iron is seen as catalyzing the formation of the hydroxyl radical or perhaps forming 

reactive iron-oxygen complexes (Smith 1989). 

Data from a study in male Long Evans rats suggested that the metabolism of hexachlorobenzene to pentachlorobenzene and other more polar metabolites proceed either through a free-radical mechanism or by 

initial formation of an arene oxide. These reactive intermediates may form covalent bonds with cellular 

constituents (such as protein amino acids or DNA nucleic acids) leading to irreversible cell damage (Lui 
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and Sweeney 1975). Several other studies have also found evidence of binding to cellular proteins by 

reactive electrophilic metabolites of hexachlorobenzene formed by cytochrome P-450 system 

(Gopalaswamy and Nair 1992; Rajamanickam and Padmanaban 1974; van Ommen et al. 1985). 

3.5.3 Animal-to-Human Extrapolations 

Studies have investigated the adverse effects of hexachlorobenzene in rats, mice, hamsters, dogs, pigs, 

and monkeys following subchronic exposure and in rats, mice, and hamsters following chronic exposure. 

Substantial bodies of both of human data and animal data are available that demonstrate qualitative 

similarities between animals and humans for such end points as porphyria and dermal lesions. Overall, 

data in animal studies do not suggest species variations in the toxicokinetics of hexachlorobenzene except 

in carcinogenic responses. The cancer toxicity data suggest that species differences exist, as 

demonstrated by multi-tumor-type responses evident in hamsters and single-tumor-type responses 

observed in mice (Cabral et al. 1977, 1979; EPA 1980b; Ertürk et al. 1986; Lambrecht et al. 1983a; Smith 

1989; Smith et al. 1985). 

3.6 TOXICITIES MEDIATED THROUGH THE NEUROENDOCRINE AXIS 

Recently, attention has focused on the potential hazardous effects of certain chemicals on the endocrine 

system because of the ability of these chemicals to mimic or block endogenous hormones. Chemicals 

with this type of activity are most commonly referred to as endocrine disruptors. However, appropriate 

terminology to describe such effects remains controversial. The terminology endocrine disruptors, 

initially used by Colborn and Clement (1992), was also used in 1996 when Congress mandated the 

Environmental Protection Agency (EPA) to develop a screening program for “...certain substances 

[which] may have an effect produced by a naturally occurring estrogen, or other such endocrine 

effect[s]...”. To meet this mandate, EPA convened a panel called the Endocrine Disruptors Screening and 

Testing Advisory Committee (EDSTAC), which in 1998 completed its deliberations and made 

recommendations to EPA concerning endocrine disruptors. In 1999, the National Academy of Sciences 

released a report that referred to these same types of chemicals as hormonally active agents. The 

terminology endocrine modulators has also been used to convey the fact that effects caused by such 

chemicals may not necessarily be adverse. Many scientists agree that chemicals with the ability to disrupt 

or modulate the endocrine system are a potential threat to the health of humans, aquatic animals, and 

wildlife. However, others think that endocrine-active chemicals do not pose a significant health risk, 

particularly in view of the fact that hormone mimics exist in the natural environment. Examples of 
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natural hormone mimics are the isoflavinoid phytoestrogens (Adlercreutz 1995; Livingston 1978; Mayr et 

al. 1992). These chemicals are derived from plants and are similar in structure and action to endogenous 

estrogen. Although the public health significance and descriptive terminology of substances capable of 

affecting the endocrine system remains controversial, scientists agree that these chemicals may affect the 

synthesis, secretion, transport, binding, action, or elimination of natural hormones in the body responsible 

for maintaining homeostasis, reproduction, development, and/or behavior (EPA 1997c). Stated 

differently, such compounds may cause toxicities that are mediated through the neuroendocrine axis. As 

a result, these chemicals may play a role in altering, for example, metabolic, sexual, immune, and 

neurobehavioral function. Such chemicals are also thought to be involved in inducing breast, testicular, 

and prostate cancers, as well as endometriosis (Berger 1994; Giwercman et al. 1993; Hoel et al. 1992). 

Hexachlorobenzene mediates multiple adverse effects through the neuroendocrine axis. Hormonal 

changes associated with exposure to hexachlorobenzene include decreased serum thyroxine 

(hypothyroidism), increased serum parathyroid hormone (hyperparathyroidism), decreased corticosterone 

released from the adrenal gland, and changes in estradiol and progesterone levels in females at certain 

times in the menstrual cycle. The female hormone changes are coincident with ovarian lesions and 

changes in female reproductive cycles in the same studies. Reduced fertility in breeding trials with hexachlorobenzene may be secondary to the ovarian effects. These alterations are described in more detail in 

Section 3.2. The effect of hexachlorobenzene on throxine appears to involve stimulation of 

dehalogenation of the hormone in the liver, rather than an effect on synthesis of the hormone in the 

thyroid (Kleiman de Pisarev et al. 1989, 1990). There is also some evidence that hexachlorobenzene may 

competitively inhibit binding of thyroxine to serum carrier proteins, further depressing circulating levels 

of the hormone (Foster et al. 1993; Van Den Berg 1990). The effects on ovarian and adrenal hormones 

have been hypothesized to reflect alterations in steroidogenesis in these tissues, possibly as a consequence 

of lipid peroxidation of mitochondrial membranes (Foster et al. 1995a, 1995b). Ultrastructural lesions 

consistent with lipid peroxidation have been observed in mitochondria from the ovaries of monkeys 

treated with hexachlorobenzene (Bourque et al. 1995); similar lesions have also been observed in rats 

(Alvarez et al. 2000). Breast cancer is another end point believed to be influenced through the 

neuroendocrine axis. Studies available to date have found little or no evidence for an association between 

hexachlorobenzene and breast cancer in humans (Dorgan et al. 1999; Guttes et al. 1998; Liljegren et al. 

1998; Mendonca et al. 1999; Moysich et al. 1998; Zheng et al. 1999). Adipose hexachlorobenzene levels 

were increased in males with undescended testis compared to controls (Hosie et al. 2000); this adverse 

effect may be related to in utero changes in hormone levels. 
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3.7 CHILDREN’S SUSCEPTIBILITY 

This section discusses potential health effects from exposures during the period from conception to 

maturity at 18 years of age in humans, when all biological systems will have fully developed. Potential 

effects on offspring resulting from exposures of parental germ cells are considered, as well as any indirect 

effects on the fetus and neonate resulting from maternal exposure during gestation and lactation. 

Relevant animal and in vitro models are also discussed. 

Children are not small adults. They differ from adults in their exposures and may differ in their 

susceptibility to hazardous chemicals. Children’s unique physiology and behavior can influence the 

extent of their exposure. Exposures of children are discussed in Section 6.6 Exposures of Children. 

Children sometimes differ from adults in their susceptibility to hazardous chemicals, but whether there is 

a difference depends on the chemical (Guzelian et al. 1992; NRC 1993). Children may be more or less 

susceptible than adults to health effects, and the relationship may change with developmental age 

(Guzelian et al. 1992; NRC 1993). Vulnerability often depends on developmental stage. There are 

critical periods of structural and functional development during both prenatal and postnatal life and a 

particular structure or function will be most sensitive to disruption during its critical period(s). Damage 

may not be evident until a later stage of development. There are often differences in pharmacokinetics 

and metabolism between children and adults. For example, absorption may be different in neonates 

because of the immaturity of their gastrointestinal tract and their larger skin surface area in proportion to 

body weight (Morselli et al. 1980; NRC 1993); the gastrointestinal absorption of lead is greatest in infants 

and young children (Ziegler et al. 1978). Distribution of xenobiotics may be different; for example, 

infants have a larger proportion of their bodies as extracellular water and their brains and livers are 

proportionately larger (Altman and Dittmer 1974; Fomon 1966; Fomon et al. 1982; Owen and Brozek 

1966; Widdowson and Dickerson 1964). The infant also has an immature blood-brain barrier (Adinolfi 

1985; Johanson 1980) and probably an immature blood-testis barrier (Setchell and Waites 1975). Many 

xenobiotic metabolizing enzymes have distinctive developmental patterns. At various stages of growth 

and development, levels of particular enzymes may be higher or lower than those of adults, and 

sometimes unique enzymes may exist at particular developmental stages (Komori et al. 1990; Leeder and 

Kearns 1997; NRC 1993; Vieira et al. 1996). Whether differences in xenobiotic metabolism make the 

child more or less susceptible also depends on whether the relevant enzymes are involved in activation of 

the parent compound to its toxic form or in detoxification. There may also be differences in excretion, 

particularly in newborns who all have a low glomerular filtration rate and have not developed efficient 
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tubular secretion and resorption capacities (Altman and Dittmer 1974; NRC 1993; West et al. 1948). 

Children and adults may differ in their capacity to repair damage from chemical insults. Children also 

have a longer remaining lifetime in which to express damage from chemicals; this potential is particularly 

relevant to cancer. 

Certain characteristics of the developing human may increase exposure or susceptibility, whereas others 

may decrease susceptibility to the same chemical. For example, although infants breathe more air per 

kilogram of body weight than adults breathe, this difference might be somewhat counterbalanced by their 

alveoli being less developed, which results in a disproportionately smaller surface area for alveolar 

absorption (NRC 1993). 

Infants and young children appeared to be especially sensitive to the effects of hexachlorobenzene in the 

Turkish grain poisoning epidemic. During the epidemic, there was an extremely high rate of mortality 

(close to 100% in some villages) in breast fed infants (under 2 years of age) of mothers known to have 

ingested the contaminated bread (Gocmen et al. 1989; Peters et al. 1982). This is in contrast to a 10% rate 

of mortality in exposed adults (Peters et al. 1982, 1987). Poisoned infants displayed a condition known as 

pembe yara or "pink sore" because of the associated skin lesions (blistering and epidermolysis and 

annular erythema) (Cripps et al. 1984; Peters et al. 1982, 1987). The infant deaths were caused by 

respiratory and cardiovascular failure resulting from the disease, and sometimes followed tremors and 

convulsions (Peters et al. 1982). Pink sore was not seen in exposed adults. Infants in this study were 

likely exposed in utero via transplacental transfer and postnatally by lactational transfer (Cripps et al. 

1984; Peters et al. 1982, 1987). No quantitation of exposure (dose and duration) was presented in any of 

these clinical reports. However, an estimated dose of 0.05–0.2 g/day (0.7–2.9 mg/kg/day for a 70-kg 

person) is considered to be reliable by the original investigators of the Turkey epidemic (Cam and 

Nigogosyan 1963). Analysis of human milk from exposed women and unexposed controls in this 

epidemic showed hexachlorobenzene concentrations of 0.51 and 0.07 ppm, respectively (Cripps et al. 

1984; Gocmen et al. 1989; Peters et al. 1982, 1987). 

Based on 20–30-year follow-up studies (Cripps et al. 1984; Peters et al. 1982), patients who were young 

children (average age 7 years) during this exposure later developed numerous dermatologic, neurologic, 

and orthopedic abnormalities associated with the developmental toxicity of hexachlorobenzene. The 

reproductive histories of 42 females exposed to hexachlorobenzene as children or young adults were also 

studied. Of the 188 pregnancies in the 42 women that occurred in a 4-year period (1977–1981), there 

were 15 fetal deaths (13 miscarriages and 2 stillbirths) and 173 live births (Peters et al. 1982, 1987); 
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however, the relevance of this study is limited because the numbers of expected miscarriages and 

stillbirths were not provided. These mothers had 0.51 ppm hexachlorobenzene in their breast milk 

compared to 0.07 ppm in unexposed controls (Gocmen et al. 1989). The fetal mortalities may be related 

to the mobilization of hexachlorobenzene from the maternal fat pool and its subsequent exposure to the 

fetus through the placenta. 

Based on partial evaluation of 63 of a planned 100 cases, Sala et al. (1999b) published a preliminary 

study, reporting that a significant association between prenatal hexachlorobenzene exposure and impaired 

development of locomotor skills had been detected in newborn babies in Flix, compared with those of 

nearby villages. A study of a less-exposed population in New York was unable to correlate hexachlorobenzene levels in umbilical blood or breast milk with infant intelligence test results (Darvill et al. 2000). 

A German case-control study found that adipose hexachlorobenzene levels in 18 male patients who 

underwent orchidopexy to correct unilateral or bilateral undescended testis (mean age 4.2 years) were 

3-fold higher compared to a group of 30 male control patients (mean age 3.5 years); this difference was 

highly significant (Hosie et al. 2000). A similar correlation was also observed for heptachloroepoxide 

(HCE), but not for other organochlorines. The weaknesses of this study are the small study size, the lack 

of age-adjustment between groups, and the potentially confounding effect of HCE. 

Although there is only limited direct evidence that hexachlorobenzene crosses the placenta in humans 

(Ando et al. 1985), animal studies have shown that hexachlorobenzene crosses the placenta readily and 

accumulates in fetal tissues (Courtney and Andrews 1985; Courtney et al. 1979; Cripps 1990; Villeneuve 

and Hierlihy 1975; Villeneuve et al. 1974a). While numerous human studies have demonstrated the 

presence of hexachlorobenzene in breast milk, animal studies have shown in addition that hexachlorobenzene concentrates in the breast milk and is transferred to the suckling neonate in considerable amounts 

(Bailey et al. 1980; Cripps 1990; Goldey et al. 1990). 

Animal studies have also confirmed that the developing organism is an especially sensitive target for 

hexachlorobenzene. Findings from laboratory animal single- and multi-generation reproductive toxicity 

studies conducted in rats exposed to hexachlorobenzene indicate that fertility, gestational viability, and 

lactational indices may be affected by hexachlorobenzene exposure (Grant et al. 1977; Kitchin et al. 

1982). Studies on prenatally exposed animals have shown immune and neurological effects at lower 

doses in the young developing animals than in adults (Goldey and Taylor 1992; Vos et al. 1979a). The 
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most sensitive end point in any study, and the basis for the chronic MRL, was liver lesions that developed 

during adulthood in rats treated with combined pre- and postnatal lifetime exposure (Arnold et al. 1985). 

Animal studies have also shown that hexachlorobenzene mediates toxicity through the neuroendocrine 

axis, with multiple effects on the thyroid gland (hypothyroidism), parathyroid gland 

(hyperparathyroidism), adrenal gland, and ovaries (Alvarez et al. 2000; Andrews et al. 1988, 1990; Den 

Besten et al. 1993; Den Tonkelaar et al. 1978; Foster et al. 1992b, 1993, 1995a, 1995b; Kimbrough and 

Linder 1974; Kleiman de Pisarev et al. 1989, 1990). Because the hormones produced by these endocrine 

organs play a crucial role in growth and development of the organism, it is not surprising that hexachlorobenzene interferes with these processes. Neuroendocrine end points have not been studied in developing 

organisms. 

3.8 BIOMARKERS OF EXPOSURE AND EFFECT 

Biomarkers are broadly defined as indicators signaling events in biologic systems or samples. They have 

been classified as markers of exposure, markers of effect, and markers of susceptibility (NAS/NRC 1989). 

Due to a nascent understanding of the use and interpretation of biomarkers, implementation of biomarkers 

as tools of exposure in the general population is very limited. A biomarker of exposure is a xenobiotic 

substance or its metabolite(s) or the product of an interaction between a xenobiotic agent and some target 

molecule(s) or cell(s) that is measured within a compartment of an organism (NAS/NRC 1989). The 

preferred biomarkers of exposure are generally the substance itself or substance-specific metabolites in 

readily obtainable body fluid(s), or excreta. However, several factors can confound the use and 

interpretation of biomarkers of exposure. The body burden of a substance may be the result of exposures 

from more than one source. The substance being measured may be a metabolite of another xenobiotic 

substance (e.g., high urinary levels of phenol can result from exposure to several different aromatic 

compounds). Depending on the properties of the substance (e.g., biologic half-life) and environmental 

conditions (e.g., duration and route of exposure), the substance and all of its metabolites may have left the 

body by the time samples can be taken. It may be difficult to identify individuals exposed to hazardous 

substances that are commonly found in body tissues and fluids (e.g., essential mineral nutrients such as 

copper, zinc, and selenium). Biomarkers of exposure to hexachlorobenzene are discussed in 

Section 3.8.1. 
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Biomarkers of effect are defined as any measurable biochemical, physiologic, or other alteration within an 

organism that, depending on magnitude, can be recognized as an established or potential health 

impairment or disease (NAS/NRC 1989). This definition encompasses biochemical or cellular signals of 

tissue dysfunction (e.g., increased liver enzyme activity or pathologic changes in female genital epithelial 

cells), as well as physiologic signs of dysfunction such as increased blood pressure or decreased lung 

capacity. Note that these markers are not often substance specific. They also may not be directly 

adverse, but can indicate potential health impairment (e.g., DNA adducts). Biomarkers of effects caused 

by hexachlorobenzene are discussed in Section 3.8.2. 

A biomarker of susceptibility is an indicator of an inherent or acquired limitation of an organism's ability 

to respond to the challenge of exposure to a specific xenobiotic substance. It can be an intrinsic genetic 

or other characteristic or a preexisting disease that results in an increase in absorbed dose, a decrease in 

the biologically effective dose, or a target tissue response. If biomarkers of susceptibility exist, they are 

discussed in Section 3.10 “Populations That Are Unusually Susceptible”. 

3.8.1 Biomarkers Used to Identify or Quantify Exposure to Hexachlorobenzene 

Human tissues and body fluids that have been analyzed for hexachlorobenzene to identify and quantify 

exposure include blood and serum (Ataniyazova et al. 2001; Glynn et al. 2000; Hagmar et al. 2001; 

Karmaus et al. 2001; Rutten et al. 1988; Sala et al. 1999b, 2001b; Schlummer et al. 1998; Siyali 1972; 

Waliszewski et al. 2001; Weiderpass et al. 2000; and many others), liver (Dewailly et al. 1999; Weistrand 

and Noren 1998), bone marrow (Bucholski et al. 1996; Scheele et al. 1995), brain (Dewailly et al. 1999), 

fat (Ansari et al. 1986; Dewailly et al. 1999; Lordo et al. 1996; Robinson et al. 1990; Scheele et al. 1995; 

Siyali 1972; Szymczynski and Waliszewski 1981; Teufel et al. 1990; Weistrand and Noren 1998), semen 

(Szymczynski and Waliszewski 1981), the placenta (Poli et al. 1999), the umbilical cord (Burse et al. 

2000; Darvill et al. 2000), and breast milk (Ataniyazova et al. 2001; Craan and Haines 1998; Czaja et al. 

1997; Darvill et al. 2000; Dewailly et al. 2000; Fitzgerald et al. 2001; Gladen et al. 1999; Gocmen et al. 

1989; Huang et al. 1989; Lunden and Noren 1998; Newsome and Ryan 1999; Scheele et al. 1995; 

Weisenberg 1986; Wickstrom et al. 1983; and others). 

Reliable methods are also available to measure hexachlorobenzene in feces (Albro and Thomas 1974; 

Koss and Koransky 1975; Schlummer et al. 1998) and urine. Trace amounts of unchanged hexachlorobenzene have been detected in urine; however, urinary metabolites are more easily detected and 

quantified (Ingebrigtsen et al. 1981, 1986; Koss et al. 1976; Lui and Sweeney 1975; Rozman et al. 1977a; 
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To-Figueras et al. 1992; van Ommen et al. 1985; Yang et al. 1978) as biomarkers for hexachlorobenzene 

exposure. Although urinary PCP and tissue hexachlorobenzene correlated in 60 patients studied, it is 

possible that the urinary PCP originated from other chlorinated hydrocarbons such as pentachlorobenzene, alpha-hexachlorocyclohexane, or pentachloronitrobenzene (Burton and Bennett 1987; Currier et 

al. 1980; Koss et al. 1986; To-Figueras et al. 1992). 

Indirect biomarkers of hexachlorobenzene exposure include measurement of gamma-glutamyl transferase 

in blood, uroporphyrin and d-ALA in urine, and coproporphyrin in feces (Koss et al. 1986; To-Figueras et 

al. 1992). Because these biomarkers are not specific for hexachlorobenzene, their usefulness in 

monitoring exposed populations is limited. 

Several studies have correlated hexachlorobenzene levels with different end points. In humans, hexachlorobenzene levels are correlated between feces and serum (To-Figueras et al. 2000), maternal and 

umbilical cord blood levels (Ataniyazova et al. 2001; Sala et al. 2001a; Waliszewski et al. 2000), length 

of breast-feeding and infant serum levels (Abraham et al. 2000), and the presence of other 

organochlorines in serum (Burse et al. 2000; Glynn et al. 2000; Hoppin et al. 2000; and others). 

Sufficient data of air levels of hexachlorobenzene have not been available to determine quantitative 

biomarkers of inhalation exposure. However, hexachlorobenzene levels have been assayed in people of 

Flix, Spain, who were exposed to hexachlorobenzene from an nearby electrochemical plant that produced 

organochlorines (Ballester et al. 2000; Grimalt et al. 1994; Herrero et al. 1999; Sala et al. 1999a, 1999b; 

To-Figueras et al. 1997). These studies found higher serum levels in factory workers compared to 

nonworkers, in male workers compared to females (presumably due to increased work-related exposure), 

in nonworkers who lived with factory workers compared to nonworkers who did not live with factory 

workers, and in people living near the factory compared to people living further away. 

3.8.2 Biomarkers Used to Characterize Effects Caused by Hexachlorobenzene 

Although not specific to hexachlorobenzene, porphyria is the primary biomarker of effect from human 

exposure to hexachlorobenzene. Disturbance of the heme biosynthesis pathway of the body's porphyrin 

metabolism in the liver is the major action of hexachlorobenzene in short- or long-term exposure. Due to 

this disturbance, abnormal levels of porphyrin precursors are found in exposed individuals. In some 

cases, porphyria cutanea tarda, displayed as scarring or cutaneous annular erythema (a condition termed 

pembe yara, or pink sore), is present. Such exposed people also exhibited painless arthritis, osteoporosis, 
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and small distinctive hands (Cripps et al. 1984; Peters et al. 1982, 1987). Increases in serum gammaglutamyl transferase, uroporphyrin (red-tinged urine), and d-ALA in the urine, and uroporphyrin and 

coproporphyrin in the stool are also indicative of the effect of hexachlorobenzene. While low levels of 

hexachlorobenzene have been found in human tissues and body fluids, such reported low levels have not 

generally been associated with adverse health effects (Booth and McDowell 1975). Recently, 

associations have been found between increased hexachlorobenzene levels and decreased interferon-γ

(Daniel et al. 2001), decreased lymphocyte IL-10 secretion (Belles-Isles et al. 2000), ear infections in 

infants (Dewailly et al. 2000), undescended testis (Hosie et al. 2000), and locomotor skill impairment in 

newborns (Sala et al. 1999b). 

For more information on biomarkers for renal and hepatic effects of chemicals see Agency for Toxic 

Substances and Disease Registry/Centers for Disease Control and Prevention Subcommittee Report on 

Biological Indicators of Organ Damage (1990) and for information on biomarkers for neurological effects 

see OTA (1990). 

3.9 INTERACTIONS WITH OTHER CHEMICALS 

Certain chemicals may interfere with the toxicity of hexachlorobenzene indirectly by influencing its 

metabolism through their actions on drug metabolizing enzymes. The duration and intensity of action of 

hexachlorobenzene are largely determined by the speed at which it is metabolized in the body by the liver 

microsomal cytochrome P-450 system. More than 200 drugs, insecticides, and other chemicals are 

known to induce the activity of liver microsomal drug-metabolizing enzymes. The characteristic 

biological actions of these chemicals are highly varied. Although there is no clear relationship between 

their actions and structures and their ability to induce enzymes, most of the inducers are lipid soluble at 

physiological pH. These inducers of the P-450 system include the following classes of drugs: hypnotics 

and sedatives (barbiturates, ethanol); anesthetic gases (methoxyflurane, halothane); central nervous 

system stimulants (amphetamine); anticonvulsants (diphenylhydantoin); tranquilizers (meprobamate); 

antipsychotics (triflupromazine); hypoglycemic agents (carbutamide); anti-inflammatory agents 

(phenylbutazone); muscle relaxants (orphenadrine); analgesics (aspirin, morphine); antihistaminics 

(diphenhydramine); alkaloids (nicotine); insecticides (chlordane, DDT, BHC, aldrin, dieldrin, 

heptachlorepoxide, pyrethrins); steroid hormones (testosterone, progesterone, cortisone); and 

carcinogenic polycyclic aromatic hydrocarbons (3-methylcholanthrene, 3,4-benzpyrene) (Klaassen et al. 

1995; Williams and Burson 1985). 
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Hexachlorobenzene has been reported to increase the activity of aryl hydrocarbon hydroxylase and other 

enzymes associated with both the 3-MC and phenobarbital-inducible isozymes of cytochrome P-450 in 

the rat (Goldstein et al. 1986). Thus, exposure to any of these enzyme inducers prior to, concurrent with, 

or after exposure to hexachlorobenzene may result in accelerated biotransformation to the more toxic 

PCP. The extent of toxicity mediated by this phenomenon is dependent on how rapidly the PCP is 

hydrolyzed or conjugated to the less chlorinated benzenes which are much less toxic. In animal studies, 

pretreatment of rats with 3-methylcholanthrene or phenobarbital increased the metabolism and toxicity of 

hexachlorobenzene (Gopalaswamy and Nair 1992; Rajamanickam and Padmanaban 1974; van Ommen et 

al. 1985; Vos et al. 1988). PCP (500 ppm) accelerated the onset of hexachlorobenzene-induced 

porphyria, indicated by an increase in urinary excretion of uroporphyrin and a decrease of porphyrin with 

2 and 3 carboxylic groups in female rats fed diets containing 1,000 ppm hexachlorobenzene (Debets et al. 

1980a). This increase occurred 3 weeks earlier in the hexachlorobenzene plus PCP-treated animals than 

in animals treated with hexachlorobenzene alone. Intraperitoneal pretreatment with diethylstilbestrol 

followed by oral administration of hexachlorobenzene to male and female rats stimulated the excretion of 

hexachlorobenzene metabolites via urine and feces (Rizzardini and Smith 1982). A 2-fold increase in the 

accumulation of hexachlorobenzene in tissues also occurred (Villeneuve et al. 1977). An exacerbation in 

the increase of hepatic accumulation and urinary excretion of uroporphyrin occurred in rats given doses of 

hexachlorobenzene (25 mg/kg/day for 12 consecutive days) when hexachlorobenzene was coadministered with methyl isobutyl ketone. The authors speculated on the involvement of hepatic isozyme 

inhibition and porphyria induction by methyl isobutyl ketone in hexachlorobenzene porphyrinogenic 

action (Krishnan et al. 1992). 

Similarly, prior or concurrent exposure to hexachlorobenzene and MFO enzyme-inhibiting substances 

(e.g., carbon monoxide; ethylisocyanide; SKF 525A, halogenated alkanes, such as CCl4; alkenes, such as 

vinyl chloride; and allelic and acetylenic derivatives) may decrease the toxicity of hexachlorobenzene by 

decreasing the rate of the hydrolytic dealkylation and hydrolysis of both parent hexachlorobenzene 

(Williams and Burson 1985). Rats treated with hexachlorobenzene in combination with the cytochrome 

P450IIIA1 (CYP3A1) inhibitor, TAO, showed a marked reduction in hexachlorobenzene-induced 

immunomodulatory effects. These results suggest that the oxidative metabolites, PCP and TCHQ, are not 

likely to be implicated in the immunostimulatory effects of hexachlorobenzene (Schielen et al. 1995a). 

Similar conclusions were reached by the investigators of a 13-week rat study to assess the role of 

oxidative metabolism in hexachlorobenzene-induced porphyria and thyroid hormone homeostasis (Den 

Besten et al. 1993) as well as in other animal studies in which pretreatment of rats with TAO decreased 
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the metabolism of hexachlorobenzene (Gopalaswamy and Nair 1992; Rajamanickam and Padmanaban 

1974; van Ommen et al. 1985; Vos et al. 1988). 

Food deprivation has also been shown to increase susceptibility of animals to the toxicity of hexachlorobenzene. During 4 weeks of exposure to hexachlorobenzene at 40 and 200 mg/kg, food deprivation 

(50%) increased the ability of hexachlorobenzene to cause liver hypertrophy and induce microsomal 

enzyme activity. A 2-fold increase in the accumulation of hexachlorobenzene in tissues also occurred 

(Villeneuve et al. 1977). Since absorption was not measured, it is not clear whether these observations 

are due to metabolic changes in fat metabolism and release of hexachlorobenzene to target organs or to 

increased fractional absorption of hexachlorobenzene. These findings were validated by Kishima et al. 

(2000); 50 ppm hexachlorobenzene did not induce liver toxicity in liver-initiated Wistar rats when 

administered in a normal diet, but caused liver damage (decreased weight, foci of altered enzyme 

expression, hypertrophy) when administered in an energy-restricted diet which provided only 50% of the 

calories in the normal diet. 

Results from other studies indicate that hexachlorobenzene has the potential to alter the toxicity of other 

chemicals. Co-treatment of hexachlorobenzene (400 μmol/kg) and 2,3,7,8-TCDD (10 or 30 μg/kg) in rats 

exacerbated both body weight loss and thymic atrophy caused by 2,3,7,8-TCDD, while hexachlorobenzene administered at doses as high as 3,000 μmol/kg did not cause any significant effects on these 

parameters (Li et al. 1989). Exposure to 4 mg/kg/day of hexachlorobenzene from 2 weeks prior to mating 

through lactation and partially through the placenta increased the LD50 value for malathion in 

17–18-day-old Wistar rat pups by more than a factor of 2. In general, the inhibitory effect of malathion 

on cholinesterase activities was decreased by pretreatment with hexachlorobenzene. The authors 

attributed the increased resistance to intoxication by malathion and reduction of esterase inhibition in the 

pups to an increase in tissue carboxylesterase activity, presumably malathionase, and a decrease in 

malaxon formation (Mendoza and Shields 1976). In a study with female Sprague-Dawley rats 

administered single gavage doses of 400 or 600 mg/kg hexachlorobenzene in corn oil, or 10 or 

12.5 mg/kg mercuric chloride, or a combination of 400 or 600 mg/kg hexachlorobenzene and 10 or 

12.5 mg/kg mercuric chloride, hexachlorobenzene and mercuric chloride interacted additively with 

respect to lethality and endocrine, kidney, and liver toxicity. Although no deaths were reported in the 

400 mg/kg hexachlorobenzene or 10 mg/kg mercuric chloride dose group animals, one death each was 

reported in the combined 400 mg/kg hexachlorobenzene plus 10 mg/kg mercuric chloride, and 600 mg/kg 

hexachlorobenzene plus 10 mg/kg mercuric chloride dose group animals; and two animals died in the 

600 mg/kg hexachlorobenzene plus 12.5 mg/kg mercuric chloride dose group animals. Similarly, mild to 
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moderate morphological changes observed in the liver, thyroid, thymus, and bone marrow of rats exposed 

to hexachlorobenzene or hexachlorobenzene plus mercuric chloride; and in the kidneys of mercuric 

chloride- or mercuric chloride plus hexachlorobenzene-exposed rats were more severe in animals that 

received a combination of hexachlorobenzene and mercuric chloride (Lecavalier et al. 1994). The 

mechanism of these interactive effects are not known. 

Iron overload aggravates hexachlorobenzene-induced porphyria and related hepatopathology. There was 

increased porphyrin excretion in female C57BL/6J mice (strains B6-Ahb

 and B6-Ahd) pretreated with iron 

and then fed diets containing 26 mg/kg/day of hexachlorobenzene for 9, 15, or 17 weeks as compared to 

rats given hexachlorobenzene alone. This is consistent with the proposition that the sustained induction 

of either P3-450 (the mouse CYP1A2 ortholog), or P1-450 (CYP1A1), or both may be a causative factor 

in the development of this disease. Furthermore, differential induction of the P3-450 (the mouse 

CYP1A2 ortholog) and P1-450 (CYP1A1) isozymes in B6-Ahb

 responsive versus B6-Ahd nonresponsive 

mice suggests that hexachlorobenzene may act through the Ah receptor (Hahn et al. 1988). Iron overload 

also caused a significantly depressed EROD (an estimate of CYP1A1 activity), in the livers of hexachlorobenzene-fed rats for 5 or 15 weeks while PROD (an estimate of CYP2B1 activity) and BROD (an 

estimate of CYP2B1 and other P-450 isozymes activity) were depressed in female Fischer 344 rats that 

received iron-dextran solution (50 mg/mL) by subcutaneous injection for 1 week and then were 

administered dietary hexachlorobenzene at a dose of 10 mg/kg/day in corn oil for 65 weeks (Smith et al. 

1993). 

The interactive effect of hexachlorobenzene with other substances in cancer induction has also been 

studied in animals. The toxicological effects of hexachlorobenzene exposure as a consequence of varying 

the dietary levels of vitamin A were evaluated in a single-generation lifetime study. There were no 

significant differences in hematological and pathological lesions in rats fed basal diets with either 0.1 or 

10 times the vitamin A content of the control diet and animals fed similar diets which also included hexachlorobenzene (Arnold et al. 1985). Hyperplastic nodules were observed in the liver lobes of 80% of 

female Fischer 344 rats pretreated with Imferon (iron-dextran) then given dietary hexachlorobenzene at 

doses of 5 or 10 mg/kg/day in arachis oil for 65 weeks. There was a high incidence of fibrin throughout 

the liver with sinusoidal telangiectasis, and nodular peliosis hepatitis and hepatocellular necrosis. The 

study proposed a nongenotoxic mechanism for tumor induction by hexachlorobenzene, concluding that 

the formation of hepatomas and hemangiomas with elements of peliosis could be explained by the 

compensatory hyperplastic responses to hepatocellular injury or necrosis and by the simultaneous loss of 
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hepatocellular cords. The study further concluded that the accumulation of iron in the liver would 

strongly potentiate the development of hepatic tumors (Carthew and Smith 1994). 

3.10 POPULATIONS THAT ARE UNUSUALLY SUSCEPTIBLE 

A susceptible population will exhibit a different or enhanced response to hexachlorobenzene than will 

most persons exposed to the same level of hexachlorobenzene in the environment. Reasons may include 

genetic makeup, age, health and nutritional status, and exposure to other toxic substances (e.g., cigarette 

smoke). These parameters result in reduced detoxification or excretion of hexachlorobenzene, or 

compromised function of organs affected by hexachlorobenzene. Populations who are at greater risk due 

to their unusually high exposure to hexachlorobenzene are discussed in Section 6.7, Populations With 

Potentially High Exposures. 

Hexachlorobenzene has been shown to elevate porphyrin levels in humans following inhalation exposure 

(Herrero et al. 1999; Sala et al. 1999b; Seldon et al. 1999) and to cause porphyria cutanea tarda (a specific 

disease resulting from elevated porphyrin levels) following oral exposure (Cam and Nigogosyan 1963; 

Cripps et al. 1984; Dogramaci 1964; Gocmen et al. 1989; Peters et al. 1982, 1987). Studies unrelated to 

hexachlorobenzene-exposure have associated the diagnosis of porphyria cutanea tarda with infections of 

HIV and hepatitis C virus (Drobacheff et al. 1998; Egger et al. 2002; Meola and Lim 1993). It is not 

known if, or the degree to which, these diseases contribute to or exacerbate one another; however, HIV 

and hepatitis C-infected individuals may have increased susceptibility to porphyria cutanea tarda 

following hexachlorobenzene exposure. 

Case studies of hexachlorobenzene poisoning in humans indicate that young children are more sensitive 

to hexachlorobenzene intoxication. Children (average age, 7 years) who had ingested hexachlorobenzene-contaminated bread during the epidemic of hexachlorobenzene poisoning in Turkey between 

1955 and 1959 developed short stature, pinched faces, osteoporosis of bones of the hand, and painless 

arthritic changes. Some of the young children in this study were presumed to have been exposed in utero 

via transplacental transfer and postnatally by lactational transfer. The children who died were between 

the ages of 1 and 2 years, and died from a disease known as pembe yara or “pink sore” (Cripps et al. 

1984; Peters et al. 1982, 1987). Analysis of human milk from exposed women and unexposed controls in 

this epidemic showed hexachlorobenzene concentrations of 0.51 and 0.07 ppm, respectively (Cripps et al. 

1984; Gocmen et al. 1989; Peters et al. 1982, 1987). 
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In laboratory animals, reduced survival of suckling offspring of lactating mothers and fetuses of mothers 

exposed to hexachlorobenzene was also reported in several studies (Arnold et al. 1985; Grant et al. 1977; 

Kitchin et al. 1982). There is evidence that hexachlorobenzene is concentrated in milk of lactating 

monkeys exposed to hexachlorobenzene, suggesting that the risk of exposure to nursing infants may be 

greater than the risk to their mothers. Blood and tissue levels in the infants were higher than in mothers, 

and infants exhibited clinical symptoms of toxicity sooner than their mothers (Bailey et al. 1980). 

3.11 METHODS FOR REDUCING TOXIC EFFECTS 

This section will describe clinical practice and research concerning methods for reducing toxic effects of 

exposure to hexachlorobenzene. However, because some of the treatments discussed may be 

experimental and unproven, this section should not be used as a guide for treatment of exposures to hexachlorobenzene. When specific exposures have occurred, poison control centers and medical toxicologists 

should be consulted for medical advice. The following texts provide specific information about treatment 

following exposures to hexachlorobenzene: 

Tintinalli JE, Ruiz E, and Krone RL, eds. 1996. Emergency medicine. A comprehensive study. 

American College of Emergency Physicians. 4th ed. The McGraw-Hill Companies, Inc. 

Goldfrank RL, et al. 1994. Goldfrank's toxicologic emergencies. 5th ed. Appleton and Lange 5th 

Edition. 

Ellenhorn MJ and Barceloux DG. 1988. Medical toxicology: Diagnosis and treatment of human 

poisoning. New York, NY: Elsevier Science Publishing Company, Inc. 

3.11.1 Reducing Peak Absorption Following Exposure 

For inhalation exposure, management commonly includes moving the exposed individual to fresh air, and 

then monitoring for respiratory distress. Injuries to the lungs are more likely when there is severe 

respiratory irritation and persistent cough. Emergency airway support and 100% humidified 

supplemental oxygen with assisted ventilation may be needed (Dreisbach 1983; Ellenhorn and Barceloux 

1988; Haddad and Winchester 1990; Rumack and Lovejoy 1991). 

Emesis may be indicated in acute ingestion unless the patient is obtunded, comatose, or convulsing, has 

lost the gag reflex, or if hexachlorobenzene was suspended in petroleum distillates, so that there is danger 

that emesis would result in aspiration. It is most effective if initiated within 30 minutes of exposure. 

Administration of mineral oil or castor oil-based cathartics are contraindicated because they tend to 
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increase intestinal absorption of lipophilic toxicants like hexachlorobenzene. Adrenergic amines 

(decongestants, bronchodilators, or caffeine) and sympathomimetics (e.g., epinephrine) are not 

recommended because they may increase myocardial irritability and produce refractory ventricular 

arrhythmias or dysrhythmias. Cholestyramine or other agents that may bind to hexachlorobenzene may 

be useful in limiting absorption (Bryson 1989; Dreisbach 1983; Ellenhorn and Barceloux 1988; Haddad 

and Winchester 1990; Rumack and Lovejoy 1991). Gastric lavage with subsequent administration of 

activated charcoal and sorbitol cathartic have been recommended in acute management to reduce 

gastrointestinal absorption. Repeated dosing with activated charcoal or cholestyramine resin may be 

administered to enhance elimination by interrupting enterohepatic circulation, as has been demonstrated 

for chlordane and kepone (Cohn et al. 1978; Ellenhorn and Barceloux 1988). 

Decontamination is the first step in reducing dermal absorption. It is recommended that decontamination 

begin immediately after the exposure; that contaminated clothing be removed; and that the skin, hair, and 

nails be washed copiously with soap and water. Leather clothing absorbs hexachlorobenzene and should 

be discarded. The use of oils could facilitate dermal absorption and their use is not recommended. If the 

material enters the eyes, thorough rinsing with sterile physiological saline is suggested (Ellenhorn and 

Barceloux 1988; Haddad and Winchester 1990; Rumack and Lovejoy 1991). If irritation, lacrimation, or 

especially pain, swelling, and photophobia persist after 15 minutes of irrigation, then ophthalmologic 

consultation to evaluate potential ocular damage is recommended (Haddad and Winchester 1990). 

3.11.2 Reducing Body Burden 

Little can be done to reduce the body burden of absorbed hexachlorobenzene (Ellenhorn and Barceloux 

1988; Haddad and Winchester 1990; Rumack and Lovejoy 1991). Diuresis is not likely to be effective 

because of the high lipophilic nature of hexachlorobenzene. Exchange transfusion, peritoneal dialysis, 

hemodialysis, and hemoperfusion are not likely to be beneficial because of the rapidity with which hexachlorobenzene leaves the blood and locates in peripheral compartments, since this substance has an initial 

large volume of distribution. However, continued treatment with charcoal and cholestyramine may be 

useful to reduce enterohepatic recirculation (Cohn et al. 1978; Ellenhorn and Barceloux 1988), although 

no differences in fecal excretion of hexachlorobenzene were observed when cholestyramine or sesame oil 

was administered orally to rats and Rhesus monkeys following ingestion of hexachlorobenzene. When 

mineral oil was added to the diet of Rhesus monkeys treated with hexachlorobenzene, a 6–9-fold increase 

in fecal excretion of hexachlorobenzene and its metabolites was observed. Continuous administration of 

mineral oil led to increased depletion of hexachlorobenzene from both the blood and adipose tissues. 
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3. HEALTH EFFECTS 

Administration of n-hexadecane enhanced fecal excretion of hexachlorobenzene about 5-fold in rats 

(Rozman et al. 1981). The barbiturates, which have been used to control poison-induced convulsions, 

may hasten metabolism and elimination of hexachlorobenzene (Smith 1991). 

Hexachlorobenzene is known to be toxic to developing perinatal animals. The particularly potent effects 

seen in nursing infants following maternal exposure to hexachlorobenzene must be recognized when 

counseling hexachlorobenzene-exposed women of childbearing age. It is recommended that plans for 

pregnancy and contraception be included in the physician's clinical assessment, with consideration to 

specialty evaluation of residual hexachlorobenzene contamination. 

3.11.3 Interfering with the Mechanism of Action for Toxic Effects 

The most serious acute toxicological effect of exposure to chlorinated hydrocarbon pesticides appears to 

be central nervous system toxicity. These organochlorine compounds interfere with the normal flux of 

sodium and potassium ions across the neuronal membrane and disrupting the normal repolarization of 

neurons, resulting in hyperexcitation and seizures. Organochlorines also inhibit the ability of calmodulin, 

a calcium mediator in the nerves, to transport calcium ions essential for the intraneuronal release of 

neurotransmitters, thereby further inhibiting repolarization and increasing neuronal sensitivity to 

subsequent stimuli resulting in hyperexcitation and seizures (Klaassen et al. 1995). Hexachlorobenzene 

appears to produce little central nervous system toxicity at low doses. At high doses, central nervous 

system depression can dominate the clinical profile (de Matteis et al. 1961). Thus, management of 

seizures with anticonvulsants is much less likely to be needed except in acute intoxication management. 

The number of acute hexachlorobenzene-intoxication patients with seizures, however, was negligible (de 

Matteis et al. 1961). Diazepam or phenobarbital is recommended for use to reduce seizures, if seizures 

occur (Bryson 1989). 

In Turkey, in 1956, prolonged oral exposure to hexachlorobenzene was associated with an outbreak of 

acquired porphyria cutanea tarda characterized by neurological, visceral, arthritic, cutaneous, and hepatic 

symptoms. The victims also exhibited bulbous, erythematous skin lesions, which progressed to atrophy, 

hyperpigmentation, hypertricosis (increased body hair), and ulcerations. Treatment for these conditions 

was primarily supportive. Anecdotal reports from Turkey indicated that chelating agents (disodium 

ethylenediaminetetraacetic acid [EDTA] and dimercaptopropanol [BAL]) administration over 3 months 

(1.5 g daily for 5 days intravenously followed by daily oral doses of 1–2 g) successfully reduced the 

symptoms of patients with hexachlorobenzene-induced porphyria (Ellenhorn and Barceloux 1988; Peters 
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3. HEALTH EFFECTS 

1956, 1993; Peters and Cripps 1985; Peters et al. 1957, 1966, 1986b). However, this method of treating 

acute porphyria has not been validated. 

The role of iron overload in the pathogenesis of hexachlorobenzene-induced porphyria has also been 

examined based on observations that 80% of patients with porphyria cutanea tarda have increased liver 

stores of iron and increased levels of uroporphyrin 1 (Smith and de Matteis 1990; Wainstok de 

Calmanovici et al. 1991). In these patients, phlebotomy often induces disease remission and a decrease in 

urinary porphyrin 1 excretion. It remains to be seen whether iron overload plays a permissive or etiologic 

role in patients exposed to porphyria-producing toxins, and whether phlebotomy has any role in the 

treatment of these patients. 

In other reports, co-administration of S-adenosyl-L-methionine (SAM) via subcutaneous injection for the 

last 15 days of treatment with an oral dose of hexachlorobenzene (100 mg/kg/day) reversed some of the 

effects of hexachlorobenzene exposure (elevation of liver porphyrin content and liver weight). It has 

been suggested that the beneficial effects of SAM on hexachlorobenzene-induced toxicity may be related 

to effects on adenosine triphosphate availability (Cantoni et al. 1990; Cuomo et al. 1991). 

A study conducted with Osteogenic Disorder Shionogi (ODS) rats provided evidence that large doses of 

ascorbic acid may inhibit chemically-induced uroporphyria in humans. This effect seems to be mediated 

by inhibition of CYP1A2-catalyzed uroporphyrinogen oxidation. However, a significant depletion of 

hepatic ascorbic acid may be required for any beneficial effect of ascorbic acid to be observed (Sinclair et 

al. 1995). 

3.12 ADEQUACY OF THE DATABASE 

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with the 

Administrator of EPA and agencies and programs of the Public Health Service) to assess whether 

adequate information on the health effects of hexachlorobenzene is available. Where adequate 

information is not available, ATSDR, in conjunction with the National Toxicology Program (NTP), is 

required to assure the initiation of a program of research designed to determine the health effects (and 

techniques for developing methods to determine such health effects) of hexachlorobenzene. 

The following categories of possible data needs have been identified by a joint team of scientists from 

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would 
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3. HEALTH EFFECTS 

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean 

that all data needs discussed in this section must be filled. In the future, the identified data needs will be 

evaluated and prioritized, and a substance-specific research agenda will be proposed. 

3.12.1 Existing Information on Health Effects of Hexachlorobenzene 

The existing data on health effects of inhalation, oral, and dermal exposure of humans and animals to 

hexachlorobenzene are summarized in Figure 3-5. The purpose of this figure is to illustrate the existing 

information concerning the health effects of hexachlorobenzene. Each dot in the figure indicates that one 

or more studies provide information associated with that particular effect. The dot does not necessarily 

imply anything about the quality of the study or studies, nor should missing information in this figure be 

interpreted as a “data need”. A data need, as defined in ATSDR’s Decision Guide for Identifying 

Substance-Specific Data Needs Related to Toxicological Profiles (Agency for Toxic Substances and 

Disease Registry 1989), is substance-specific information necessary to conduct comprehensive public 

health assessments. Generally, ATSDR defines a data gap more broadly as any substance-specific 

information missing from the scientific literature. 

As shown in Figure 3-5, there are data available regarding a number of toxic end points in humans 

exposed to hexachlorobenzene by inhalation and ingestion, including death, chronic systemic toxicity, 

immunological, neurological, reproductive and developmental effects, and cancer. However, the extent 

of the data available is limited, particularly for inhalation exposure. No data were located regarding the 

toxicity of hexachlorobenzene by dermal exposure in humans. Hexachlorobenzene has been well studied 

in animals following oral exposure; the full range of end points have been assessed. However, only 

immunological effects have been studied in animals following inhalation exposure and no studies at all 

were located regarding toxicity of hexachlorobenzene in animals following dermal exposure. 

3.12.2 Identification of Data Needs 

Acute-Duration Exposure. No information was located regarding health effects in humans 

following acute-duration exposure to hexachlorobenzene by any route. The available animal data were 

sufficient to identify target organs of acute oral exposure. Extensive study of animals exposed by the oral 

route for acute durations has identified doses producing a wide range of health effects, including 

porphyria and other hepatic effects, renal tubular lesions, changes in thyroid and reproductive hormone 

levels, impaired male reproductive performance, developmental effects ranging from subtle 
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3. HEALTH EFFECTS 

neurobehavioral effects in neonates to fetotoxic and teratogenic effects, overt neurological effects, and 

lethality (Bouthillier et al. 1991; Courtney et al. 1976; De Matteis et al. 1961; Foster et al. 1993; Goldey 
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3. HEALTH EFFECTS 

Figure 3-5. Existing Information on Health Effects of Hexachlorobenzene 
HEXACHLOROBENZENE 175 

3. HEALTH EFFECTS 

and Taylor 1992; Goldstein et al. 1978; Kennedy and Wigfield 1990; Khera 1974a; Mehendale et al. 

1975; Simon et al. 1979). Most of the existing acute studies were conducted in rats, but mice and guinea 

pigs were also studied. Pharmacokinetic models for oral exposure in rats and humans are available 

(Freeman et al. 1989; Yesair et al. 1986), and provided good correlations between predicted yield of 

metabolites and experimentally observed liver toxicity (porphyria and increased liver weight reported by 

Koss et al. 1978 and Kuiper-Goodman et al. 1977), between estimated concentrations of hexachlorobenzene in the brain and observed central nervous system effects (tremors and other signs reported by 

Koss et al. 1978 and Kuiper-Goodman et al. 1977), and between estimated extent of lactational transfer of 

hexachlorobenzene and observed offspring mortality (data from Kitchen et al. 1982). The database was 

sufficient to support derivation of an acute oral MRL for hexachlorobenzene. Acute studies by inhalation 

exposure in animals were limited to a single study of immunological effects (Sherwood et al. 1989). No 

acute dermal studies in animals were located. Although an acute study in rats suggested that hexachlorobenzene can be absorbed across the skin (Koizumi 1991), pharmacokinetic models for inter-route 

extrapolation are not available. Additional acute studies by the inhalation and dermal routes would 

identify acute effect levels for these routes of exposure. 

Intermediate-Duration Exposure. A few case-control studies have found evidence of 

developmental toxicity in newborn humans (Belles-Isles et al. 2000; Hosie et al. 2000); no information 

was located regarding health effects in humans following intermediate-duration exposure to hexachlorobenzene by any route. Extensive study of animals exposed by the oral route for intermediate durations 

has identified doses producing a broad spectrum of health effects, including porphyria and other hepatic 

effects, renal tubular lesions, pulmonary lesions, cardiac lesions, anemia and leukocytosis, osteosclerotic 

changes in bone, necrotic lesions in muscle, skin lesions, thymic atrophy, splenomegaly and altered 

spleen morphology, lymph node histopathology, altered immunoglobulin levels, suppression of immune 

function, changes in the thyroid, parathyroid, and adrenal glands and associated hormone levels, ovarian 

and testicular lesions, alterations in female menstrual cycling and reproductive hormone levels, reduced 

fertility, developmental effects ranging from subtle neurobehavioral effects in neonates to fetotoxic 

effects and pup death, overt neurological effects, and lethality (Andrews et al. 1988, 1989, 1990; 

Babineau et al. 1991; Bourque et al. 1995; Bouthillier et al. 1991; Den Besten et al. 1993; Den Tonkelaar 

et al. 1978; Foster et al. 1992a, 1992b; Iatropoulos et al. 1976; Jarrell et al. 1993; Kennedy and Wigfield 

1990; Kimbrough and Linder 1974; Knauf and Hobson 1979; Koss et al. 1978; Kuiper-Goodman et al. 

1977; Lilienthal et al. 1996; Loose et al. 1978, 1981; Ockner and Schmid 1961; Schielen et al. 1995a, 

1995b; Smith et al. 1985; Vos et al. 1979a, 1979b, 1983; others). Most of the existing intermediateduration studies were conducted in rats, but monkeys, mice, rabbits, dogs, and pigs were also studied. 
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Pharmacokinetic models for oral exposure in rats and humans are available (Freeman et al. 1989; Yesair 

et al. 1986), and provided good correlations between predicted yield of metabolites and experimentally 

observed liver toxicity (porphyria and increased liver weight reported by Koss et al. 1978 and KuiperGoodman et al. 1977), between estimated concentrations of hexachlorobenzene in the brain and observed 

central nervous system effects (tremors and other signs reported by Koss et al. 1978 and Kuiper-Goodman 

et al. 1977), and between estimated extent of lactational transfer of hexachlorobenzene and observed 

offspring mortality (data from Kitchen et al. 1982). The database was sufficient to support derivation of 

an intermediate-duration oral MRL for hexachlorobenzene. Intermediate-duration studies by inhalation 

exposure in animals were limited to a single study of immunological effects (Sherwood et al. 1989). No 

intermediate-duration dermal studies in animals were located. Pharmacokinetic models for inter-route 

extrapolation are not available. Additional intermediate-duration studies by the inhalation and dermal 

routes would identify intermediate-duration effect levels for these routes of exposure. 

Chronic-Duration Exposure and Cancer. There are data available on humans chronically 

exposed to hexachlorobenzene by the inhalation and oral routes, but no quantitative exposure 

information. The inhalation data are very limited, but tentatively found effects on the liver and immune 

system of exposed individuals (Queiroz et al. 1997, 1998a, 1998b; Richter et al. 1994; Selden et al. 1999). 

The oral data much more clearly identified the liver, skin, bone, thyroid, and central nervous system as 

target tissues for hexachlorobenzene in chronically exposed people (Cam and Nigogosyan 1963; Cripps et 

al. 1984; Peters et al. 1982, 1987). Based on very limited data, the original investigators of the Turkey 

epidemic estimated that the daily average oral dose was 0.05–0.2 g/day (0.7–2.9 mg/kg/day for a 70-kg 

person) (Cam and Nigogosyan 1963). No information regarding chronic dermal exposure in humans was 

located. Chronic oral animal studies identified dose levels associated with systemic (cardiovascular, 

gastrointestinal, hematological, hepatic, renal, and dermal), immunological, overt neurological, and 

developmental effects, as well as death (Arnold et al. 1985; Cabral et al. 1977, 1979; Gralla et al. 1977a; 

Mollenhauer et al. 1975; Smith and Cabral 1980; Smith et al. 1985, 1989, 1993; others). The database 

was sufficient to support derivation of a chronic oral MRL for hexachlorobenzene. No chronic animal 

studies were located using inhalation or dermal exposure. Pharmacokinetic models for inter-route 

extrapolation are not available. Additional chronic studies by the inhalation and dermal routes would 

identify chronic effect levels for these routes of exposure. 

Data from people exposed to hexachlorobenzene by inhalation provide weak evidence for an association 

between exposure to hexachlorobenzene and cancer of the thyroid, brain, and liver (Grimalt et al. 1994; 

Selden et al. 1989), while very limited data from orally exposed people showed no increase in cancer risk 
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(Cripps et al. 1984; Peters et al. 1982). One case-control study associated elevated adipose levels of 

hexachlorobenzene with increased risk of breast cancer (Dewailly et al. 1994), but other case-control 

studies have not found any relationship between body burdens of hexachlorobenzene and breast cancer 

(Dorgan et al. 1999; Falck et al. 1992; Guttes et al. 1998; Liljegren et al. 1998; Mendonca et al. 1999; 

Moysich et al. 1998; Mussalo-Rauhamaa et al. 1990; Zheng et al. 1999), bone sarcoma, or leukemia 

(Hardell et al. 1997; Scheele et al. 1996). The available epidemiology reports taken together do not 

support an association between hexachlorobenzene exposure and increased cancer incidence, but their 

limitations (including small study sizes and potentially confounding effects of other organochlorines) 

preclude considering them evidence of noncarcinogenicity. Because hexachlorobenzene produces 

porphyria, it is noteworthy that several human studies have associated porphyria with increased incidence 

of liver cancer (Axelson 1986; Fracanzani et al. 2001). 

Several animal studies have demonstrated that oral exposure to hexachlorobenzene increases the 

incidence of tumor formation. The evidence of carcinogenicity is strongest in the liver; 

hexachlorobenzene has been shown to induce hyperplasia (in rats, mice, pigs, dogs, and monkeys), 

metaplasia (in rats), benign tumors (hepatoma in mice and rats; hemagniohepatoma and bile duct 

adenoma in rats), and malignant tumors (hepatocarcinoma in rats, mice, and hamsters; bile duct 

adenocarcinoma in rats) (Arnold et al. 1985; Cabral et al. 1979; Ertürk et al. 1986; Pereira et al. 1982; 

Smith and Cabral 1980; Smith et al. 1985). Additionally, exposure to hexachlorobenzene has been shown 

to induce renal metaplasia, adenomas, and renal cell carcinomas (in rats, mice, and hamsters); 

lymphosarcomas (in rats, mice, and hamsters); adrenal hyperplasia and pheochromocytoma (in rats); 

parathyroid adenomas (in rats); and hemangioendothelioma and thyroid tumors (in hamsters) (Arnold et 

al. 1985; Den Besten et al. 1993; Ertürk et al. 1986; Kimbrough and Linder 1974). No animal cancer 

bioassays by inhalation or dermal exposure were located. Pharmacokinetic models for inter-route 

extrapolation are not available. Based on these findings in animals, hexachlorobenzene is considered a 

probable human carcinogen. Additional epidemiological studies of people with known hexachlorobenzene exposure would enable better assessment of the carcinogenic risk of this chemical to humans. 

Genotoxicity. Human genotoxicity data for hexachlorobenzene are limited to a case study (route of 

exposure unknown) and in vitro studies in human cell lines. The frequency of micronuclei in peripheral 

lymphocyte was increased in 41 workers exposed to a mixture of chlorinated solvents that included hexachlorobenzene (da Silva Augusto et al. 1997). Hexachlorobenzene did not produce chromosomal 

aberrations in human peripheral lymphocytes in vitro (Siekel et al. 1991), but did produce weak positive 

results in assays for DNA fragmentation and micronuclei formation in primary cultures of human 
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hepatocytes (Canonero et al. 1997) and minimal formation of DNA adducts in cultured human Hep G2 

hepatoma cells (Dubois et al. 1997). No information on the genotoxicity of hexachlorobenzene in 

animals by inhalation or dermal exposure was located. Hexachlorobenzene did not cause gene mutations 

or unscheduled DNA repair in microbial assays (Gopalaswamy and Nair 1992; Haworth et al. 1983; 

Siekel et al. 1991), and did not produce dominant lethal mutations in orally-exposed rats (Khera 1974a; 

Simon et al. 1979), or bind strongly to rat DNA in vitro or in vivo (Gopalaswamy and Nair 1992). 

However, hexachlorobenzene did produce DNA fragmentation in cultured rat hepatocytes (Canonero et 

al. 1997) and DNA adducts in fetal hepatocytes from rats and quail (Dubois et al. 1997). Additional 

studies employing other in vivo and in vitro assays would be useful to determine the genotoxic potential 

of hexachlorobenzene. 

Reproductive Toxicity. No data were located regarding reproductive effects in humans or animals 

with inhalation or dermal exposure. Several miscarriages and stillbirths were reported among people with 

previous oral exposure to hexachlorobenzene (Peters et al. 1982, 1987), but it is not clear that the rate of 

miscarriages was significantly higher than normal for this population. One study associated elevated 

hexachlorobenzene blood levels with increased risk for spontaneous abortion (Jarrell et al. 1998), but 

other studies did not (Gerhard et al. 1998; Leoni et al. 1986, 1989). The cessation of agricultural uses of 

hexachlorobenzene in Xixin did not affect reproductive outcomes there (Huang et al. 1989). Animal 

studies using oral exposure have identified doses associated with reproductive effects, including ovarian 

lesions and hormonal and menstrual changes, in female rats and monkeys (Alvarez et al. 2000; Babineau 

et al. 1991; Bourque et al. 1995; Foster et al. 1992a, 1992b, 1993, 1995a; Iatropoulos et al. 1976; Jarrell 

et al. 1993; Muller et al. 1978; Sims et al. 1991), reduced fertility in rats (Grant et al. 1977), reduced 

mating index in male rats (Simon et al. 1979), and testicular effects in rats and pigs (including increased 

weight, degenerative lesions, and retarded maturation) (Den Tonkelaar et al. 1978; Gralla et al. 1977a; 

Smith et al. 1985). The intermediate-duration MRL for oral exposure is based on ovarian effects in 

monkeys. Reproductive effects have not been studied in animals by inhalation or dermal exposure, and 

pharmacokinetic models for inter-route extrapolation are not available. Epidemiological studies of people 

with known hexachlorobenzene exposure would be useful to establish whether these effects are also seen 

in people. Additional mechanistic studies to better understand the ovarian and hormonal changes might 

also help establish the relevance of these findings to humans. 
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Developmental Toxicity. No studies on the developmental effects of hexachlorobenzene in humans 

following dermal exposure or in laboratory animals following inhalation or dermal exposure were 

identified, and pharmacokinetic models for inter-route extrapolation are not available. Dramatic 

developmental toxicity (high mortality, skin lesions) was seen in infants whose mothers consumed bread 

contaminated with hexachlorobenzene (Cripps et al. 1984; Gocmen et al. 1989; Peters et al. 1982); this 

study clearly established hexachlorobenzene as a developmental toxicant. Other human studies have 

found suggestive evidence linking hexachlorobenzene exposure with developmental toxicity, e.g., 

locomotor skill impairment associated with inhalation exposure (Sala et al. 1999b) and increased risk of 

undescended testis (route of exposure unknown) (Hosie et al. 2000), although an additional study found 

no correlation between hexachlorobenzene levels (in blood and milk) and infant intelligence test results 

(Darvill et al. 2000). Studies in orally exposed rats have demonstrated neurodevelopmental (Goldey and 

Taylor 1992; Lilienthal et al. 1996) and immunodevelopmental effects (Barnett et al. 1987; Vos et al. 

1979a, 1983), reduced neonatal viability and growth (Grant et al. 1977; Kitchin et al. 1982; Vos et al. 

1979a, 1983), and some evidence of teratogenic abnormalities (Courtney et al. 1976; Khera 1974a). 

Hexachlorobenzene caused neurological, hepatic, and cardiovascular effects, as well as death, in 

lactionally exposed Rhesus monkey infants (Bailey et al. 1980; Iatropoulos et al. 1978). Additional 

studies that included an assessment of more sensitive end points, such as endocrine changes and 

neurological or immunological effects, and an investigation of different periods of developmental 

sensitivity (such as prenatal versus postnatal exposures) would contribute to a clearer understanding of 

the developmental toxicity of hexachlorobenzene. 

Immunotoxicity. Studies on the immunotoxicity of hexachlorobenzene in humans following oral or 

dermal exposure are lacking. Occupational studies have associated inhalation exposure to 

hexachlorobenzene with effects on immunological parameters (neutrophil chemotaxis and cytolytic 

activity, serum immunoglobulin and IFN-γ levels) (Daniel et al. 2001; Queiroz et al. 1997, 1998a, 1998b; 

Richter et al. 1994). Case-control studies have associated increased body burdens of hexachlorobenzene 

(putatively resulting from consumption of contaminated food) with alterations in markers of immune 

function and susceptibility to infection (Belles-Isles et al. 2000; Dewailly et al. 2000). No studies on the 

immunotoxicity of hexachlorobenzene in animals after dermal exposure were located. An intermediateduration study found slight decreases in humoral and pulmonary cellular defenses of rats exposed to 

hexachlorobenzene via inhalation (Sherwood et al. 1989). Following oral exposure, immunosuppression 

has been observed in rats, mice, monkeys, and bears (Bernhoft et al. 2000; Carthew et al. 1990; 

Iatropoulos et al. 1976; Loose et al. 1977, 1981; Michielsen et al. 1997; Silkworth and Loose 1981; Van 

Loveren et al. 1990) and at least a partial stimulation of the immune system in rats and dogs (Gralla et al. 
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1977a; Kennedy and Wigfield 1990; Koss et al. 1978; Schielen et al. 1993, 1995b; Vos et al. 1979a, 

1979b). Additionally, a number of animal studies have observed inflammation and immune cell 

infiltration in tissues such as the liver (Arnold et al. 1985; Ertürk et al. 1986), respiratory tract (Goldstein 

et al. 1978; Kimbrough and Linder 1974; Kitchin et al. 1982; Michielsen et al. 1997, 1999, 2001; Vos et 

al. 1979a, 1983), and skin (Koss et al. 1978; Michielsen et al. 1997, 2000; Schielen et al. 1993, 1995b; 

Torinuki et al. 1981) following oral exposure to hexachlorobenzene. Additional chronic-duration studies 

in humans in the workplace would identify effect levels for immunotoxicity. 

Neurotoxicity. No information regarding neurotoxicity in humans following dermal exposure was 

located. Following ingestion of bread contaminated with hexachlorobenzene, observed neurological 

effects included profound weakness, loss of muscle control (inability to handle utensils, myotonia 

[delayed muscle relaxation after an initial contraction], and cogwheeling [irregular jerkiness of movement 

due to increased muscle tone as seen in Parkinson’s disease]), paresthesia (spontaneous tingling or 

burning sensations), and sensory shading (graded sensory loss indicative of polyneuropathy) (Cam and 

Nigogosyan 1963; Gocmen et al. 1989; Peters et al. 1982, 1987). Inhalation data are limited to suggestive 

evidence linking inhalation exposure to hexachlorobenzene with impaired development of locomotor 

skills in newborn babies (Sala et al. 1999a). A case-control study did not associate umbilical blood or 

breast milk hexachlorobenzene levels with infant intelligence test results (Darvill et al. 2000). No studies 

on the neurotoxicity of hexachlorobenzene in animals after inhalation or dermal exposure were located. 

Oral studies, primarily in rats but also in mice, rabbits, pigs, monkeys, and quail, have demonstrated 

serious neurological effects such as convulsions, tremors (intermittent and constant), lethargy, and 

progressive weakness (Cabral et al. 1979; Cripps 1990; De Matteis et al. 1961; Den Besten et al. 1993; 

Den Tonkelaar et al. 1978; Grant et al. 1977; Hahn et al. 1988; Kennedy and Wigfield 1990; Kimbrough 

and Linder 1974; Khera 1974a; Knauf and Hobson 1979; Ockner and Schmid 1961; others) as well as 

neurobehavioral effects in rats following developmental exposure (Goldey and Taylor 1992; Lilienthal et 

al. 1996; Taylor and Goldey 1990), changes in adult rat brain chemistry (Billi de Catabbi et al. 2000b; 

Cochon et al. 2001), and electrophysiological changes in the brains of adult dogs (Sufit et al. 1986; 

Sundlof et al. 1981). The acute-duration MRL for oral exposure is based on neurobehavioral changes in 

rats following developmental exposure (Goldey and Taylor 1992). Additional inhalation and dermal 

studies would identify the potential neurotoxicity of hexachlorobenzene by these routes of exposure. 
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Epidemiological and Human Dosimetry Studies. No information regarding the adverse effects 

of hexachlorobenzene in humans following dermal exposure is available, and no human dosimetry data 

are available. Health effects (death, systemic [e.g. liver, skin, bone, and thyroid], neurological, 

developmental, and endocrine) were identified in cohorts from a group of approximately 4,000 people 

orally exposed in Turkey to hexachlorobenzene (in contaminated bread) between 1955 and 1959 (Booth 

and McDowell 1975; Cam and Nigogosyan 1963; Cripps et al. 1984; Gocmen et al. 1989; Peters et al. 

1982, 1987; Selden et al. 1989). Related to inhalation exposure, health effects (systemic [hepatic, renal, 

and endocrine] and neurological) have been identified in residents of a rural town (Flix, Spain) with 

airborne hexachlorobenzene pollution attributed to a nearby organochlorine factory, workers from that 

factory, and other people with occupational exposure (Grimalt et al. 1994; Herrero et al. 1999; Queiroz et 

al. 1997, 1998a, 1998b; Richter et al. 1994; Sala et al. 1999a, 1999b, 2001a; To-Figueras et al. 1997). 

Multiple case-studies have investigated possible associations between body burdens of hexachlorobenzene levels (in blood, fat, urine, and feces) with multiple health effects (Belles-Isles et al. 2000; 

Darvill et al. 2000; Dewailly et al. 1994, 2000; Dorgan et al. 1999; Falck et al. 1992; Gerhard et al. 1998; 

Guttes et al. 1998; Hagmar et al. 2001; Hosie et al. 2000; Leoni et al. 1986, 1989; Liljegren et al. 1998; 

Mendonca et al. 1999; Moysich et al. 1998; Mussalo-Rauhamaa et al. 1990; Zheng et al. 1999; others). 

None of these human studies have provided reliable direct exposure data (dose and duration); therefore, 

no evidence of an exposure-response relationship has been possible. Further studies of populations with 

elevated exposures to hexachlorobenzene (e.g., occupational, consumption of fish from contaminated 

areas) would provide additional information useful in assessing dosimetry and health effects such as 

reproductive and developmental toxicity. 

Biomarkers of Exposure and Effect. 

Exposure. Hexachlorobenzene has been measured in human blood and serum, liver, bone marrow, brain, 

fat, semen, placenta, umbilical cord (and cord blood), breast milk, feces, and urine (Ataniyazova et al. 

2001; Bucholski et al. 1996; Burse et al. 2000; Dewailly et al. 1999; Poli et al. 1999; Schlummer et al. 

1998; Szymczynski and Waliszewski 1981; many others). Metabolites of hexachlorobenzene (including 

chlorophenols, pentachlorobenzene, alpha-hexachlorocyclohexane, or pentachloronitrobenzene) have 

been measured in blood, urine, and feces (Ingebrigtsen et al. 1981, 1986; Koss et al. 1976; Lui and 

Sweeney 1975; Rozman et al. 1977a; To-Figueras et al. 1992; van Ommen et al. 1985; Yang et al. 1978). 

Indirect biomarkers used to detect intermediate- and chronic-exposure to hexachlorobenzene exposure 

include measurement of gamma-glutamyl transferase in blood, uroporphyrin and d-ALA in urine, and 

coproporphyrin in feces (Koss et al. 1986; To-Figueras et al. 1992); because of their lack of specificity, 
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the usefulness of these biomarkers is limited. Further studies to develop additional, more sensitive 

biomarkers of exposure that are specific for hexachlorobenzene exposure would be especially useful in 

the monitoring of people living near hazardous waste sites. 

Effect. Porphyria is the primary biomarker of effect from human acute, intermediate, and chronic 

exposure to hexachlorobenzene. Studies of an orally exposed population have diagnosed several unusual 

disease states of porphyria cutanea tarda, including dermal lesions (pembe yara or “pink sore” and kara 

yara or “black sore,” associated with photosensitivity, dermal fragility and scarring, hyperpigmentation 

and hirsutism) and small distinctive hands (shortened and spindled fingers with painless swelling and 

osteoporosis) (Cripps et al. 1984; Gocmen et al. 1989; Peters et al. 1982, 1987). Increases in serum 

gamma-glutamyl transferase, uroporphyrin and d-ALA in the urine (red-tinged urine), and uroporphyrin 

and coproporphyrin in the stool are also indicative of the effect of hexachlorobenzene (Booth and 

McDowell 1975; others). Additional studies that identified alternative biomarkers would complement 

these existing biomarkers. Moreover, direct assessments of exposure would facilitate the identification of 

effect levels. 

Absorption, Distribution, Metabolism, and Excretion. One study was located in which 

gastrointestinal absorption of hexachlorobenzene in humans was quantified (Schlummer et al. 1998). 

Information regarding absorption in humans following inhalation exposure is based on observations of 

toxicity (Grimalt et al. 1994; Herrero et al. 1999; Queiroz et al. 1997, 1998a, 1998b; Richter et al. 1994; 

Sala et al. 1999b; Selden et al. 1997; To-Figueras et al. 1997). No experimental information regarding 

absorption in humans by dermal exposures was located; dermal data from a rat study were used to 

develop a compartment model for application to a 70-kg worker (Koizumi 1991). Only one dermal 

absorption study (Koizumi 1991) and no inhalation absorption studies in animals are available. Animal 

data suggest that oral absorption is rapid if dissolved in a lipid, but absorption from aqueous solution is 

low (Ingebrigtsen and Nafstad 1983; Koss and Koransky 1975). Additional information on the 

absorption of hexachlorobenzene, especially by the inhalation route, as well as data regarding 

enterohepatic circulation and gastrointestinal reabsorption would allow more accurate estimations of 

exposure and evaluation of route-specific differences in hexachlorobenzene toxicity. 

No data on distribution in humans following inhalation or dermal exposure or in animals following 

inhalation exposure were available, but limited information on the distribution of hexachlorobenzene 

following oral exposure was located. Available data suggest that hexachlorobenzene is preferentially and 

rapidly distributed to tissues with high lipid content (Cripps 1990; Ingebrigtsen and Nafstad 1983; Jarrell 
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3. HEALTH EFFECTS 

et al. 1993; Mehendale et al. 1975; Verschueren 1983). Additional distribution studies would provide 

information regarding the tissue doses associated with adverse effects. 

The metabolism of hexachlorobenzene has not been studied in humans. Studies in monkeys and rats 

indicate that hexachlorobenzene is metabolized to less chlorinated benzenes, chlorinated phenols, other 

minor metabolites, and glucuronide and glutathione conjugates (Ingebrigtsen and Nafstad 1983; 

Ingebrigtsen et al. 1981, 1986; Jansson and Bergman 1978; Koss et al. 1986; Rozman et al. 1977a). 

Because differences in metabolism may occur with differences in the route of exposure, it would be 

useful to have more data on inhalation and dermal metabolic studies as a comparison with the available 

oral studies. 

No studies were located regarding excretion of hexachlorobenzene in animals or humans following 

inhalation or dermal exposure. Oral studies in animals indicate that the parent hexachlorobenzene is 

excreted primarily in feces, while metabolites were detected in urine (Albro and Thomas 1974; 

Ingebrigtsen et al. 1981; Mehendale et al. 1975; Rozman et al. 1977a, 1981; To-Figueras et al. 1991). 

Studies on excretion following inhalation and dermal exposure to hexachlorobenznene would be useful to 

determine if excretion patterns vary with different routes. 

PBPK models for hexachlorobenzene have been developed by Yesair et al. (1986) and Freeman et al. 

(1989). The Yesair et al. (1986) model describes the absorption, distribution, and elimination of ingested 

hexachlorobenzene in growing rats and humans. The Freeman et al. (1989) model describes distribution 

and excretion of intravenously injected hexachlorobenzene in growing rats. Additional PBPK models to 

extrapolate from high- to low-exposure and between routes of exposure would aid in risk analysis. 

Comparative Toxicokinetics. Although no toxicokinetic information is available for humans 

following dermal or inhalation exposure, data from a rat study were used to develop a compartment 

model for application to a 70-kg worker (Koizumi 1991). 

Overall, data in animal studies do not indicate the toxicokinetics of hexachlorobenzene are similar among 

species (Albro and Thomas 1974; Cripps 1990; Goldey et al. 1990; Iatropoulos et al. 1975; Koizumi 

1991; Koss et al. 1986; Rozman et al. 1977a; others). Differences observed in absorption may be related 

to vehicle and route of administration (Iatropoulos et al. 1975; Ingebrigtsen and Nafstad 1983; Koss and 

Koransky 1975; Lecavalier et al. 1994; Sundlof et al. 1982; others), and no remarkable differences have 
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been seen for distribution, metabolism, or excretion (Cripps 1990; Ingebrigtsen and Nafstad 1983; Jarrell 

et al. 1993; Mehendale et al. 1975; Verschueren 1983; others). 

PBPK models for hexachlorobenzene have been developed for rats to describe the absorption, 

distribution, and elimination of ingested hexachlorobenzene, but were not considered appropriate for 

inter-species extrapolations (Freeman et al. 1989; Yesair et al. 1986). One of the models was adapted for 

human modeling by using the same model structure with human physiological parameter values (Yesair et 

al. 1986). Although no empirical data are available on the dermal absorption of hexachlorobenzene in 

humans, dermal data from a rat study were used to develop a compartment model for application to a 

70-kg worker (Koizumi 1991). 

Further development of a human toxicokinetic model would be valuable in assessing risks to human 

health from hexachlorobenzene exposure. 

Methods for Reducing Toxic Effects. Although available poison-treatment recommendations 

provide some guidance for reducing the toxic effects of absorbed hexachlorobenzene through inhalation, 

oral, or dermal exposures, these recommendations are not specific to hexachlorobenzene. 

Recommendations to reduce peak absorption following exposure include such general procedures as 

moving the individual to fresh air following inhalation exposure, emesis and gastric lavage with activated 

charcoal following oral exposure, and removal of contaminated clothing and washing of the skin 

following dermal exposure. Little can be done to reduce body burden of hexachlorobenzene. Treatments 

that interfere with the mechanism of action for toxicity or repair tissue damage have not been developed 

specifically for hexachlorobenzene. However, some general recommendations are available, such as use 

of diazepam or phenobarbital to control convulsions related to hexachlorobenzene exposure. 

Development of methods for reducing toxic effects targeted specifically to hexachlorobenzene would be 

useful, and additional studies into the mechanisms of action would support this goal. 

Children’s Susceptibility. No data regarding children’s susceptibility following dermal or 

inhalation exposure to hexachlorobenzene were identified. The available human data suggest that infants 

and young children are at increased risk from exposure to hexachlorobenzene compared to adults (Cripps 

et al. 1984; Gocmen et al. 1989; Peters et al. 1982, 1987). Studies of an orally exposed population 

reported 95% mortality in exposed infants (under 2 years of age) associated with dermal lesions; 

adolescents (between the ages of 6 and 15 years) exhibited health effects (including 10% mortality and 

dermal lesions) more frequently than adults (Cripps et al. 1984; Gocmen et al. 1989; Peters et al. 1982, 
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3. HEALTH EFFECTS 

1987). Other studies focused on children’s health found suggestive evidence of neurological and 

immunological effects, but did not assess exposure (Belles-Isles et al. 2000; Darvill et al. 2000; Dewailly 

et al. 2000; Hosie et al. 2000; Sala et al. 1999). Although immunological effects have been seen in 

humans exposed as adults (Richter et al. 1994; Queiroz et al. 1997, 1998a, 1998b), neurological effects 

have not; therefore, children may be more susceptible than adults to the neurotoxicity of 

hexachlorobenzene. 

No animal studies relevant to children’s susceptibility following dermal or inhalation exposure were 

identified. Animal studies have confirmed that hexachlorobenzene is transferred to the developing 

organism through the placenta in utero and via lactation after birth, and that the developing animals 

exhibited signs of toxicity (such as reduced survival and anatomical abnormalities) not seen in parental 

animals at the same exposure levels (Arnold et al. 1985; Bailey et al. 1980; Courtney et al. 1979; Grant et 

al. 1977; Iatropoulos et al. 1978; Khera 1974a; Kitchin et al. 1982; others). These experiments suggest 

that the ability of hexachlorobenzene to sensitively affect the developing organism may be related to its 

demonstrated capabilities to mediate toxicity through the neuroendocrine axis (Alvarez et al. 2000; 

Andrews et al. 1988, 1990; Den Besten et al. 1993; Den Tonkelaar et al. 1978; Foster et al. 1992b, 1993, 

1995a, 1995b; Kimbrough and Linder 1974; Kleiman de Pisarev et al. 1989, 1990). However, endocrine 

end points have not been monitored in developing organisms. A developmental study that included 

assessment of endocrine end points along with other sensitive end points such as neurobehavioral and 

immune function would be useful to determine the role of endocrine changes with regard to these effects, 

and would identify critical levels of effect. 

No data were located concerning whether pharmacokinetics of hexachlorobenzene in children are 

different from adults. A PBPK model (Yesair et al. 1986) has modeled fetal and breast milk 

compartments, in humans as well as rats, for oral exposure to hexachlorobenzene. Two PBPK models 

(Freeman et al. 1989; Yesair et al. 1986) have characterized the pharmacokinetics of hexachlorobenzene 

in growing rats, and have been validated using experimental data. No information regarding biomarkers 

of exposure and effect or potential interactions of hexachlorobenzene with other chemicals pertinent to 

children’s susceptibility were identified. There are no pediatric-specific methods to reduce peak 

absorption of hexachlorobenzene following exposure, or to reduce body burden, or to interfere with 

mechanisms of action for hexachlorobenzene. 

Data needs relating to both prenatal and childhood exposures, and developmental effects expressed either 

prenatally or during childhood, are discussed in detail in the Developmental Toxicity subsection above. 
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3. HEALTH EFFECTS 

Child health data needs relating to exposure are discussed in 6.8.1 Identification of Data Needs: 

Exposures of Children. 

3.12.3 Ongoing Studies 

Ongoing studies pertaining to hexachlorobenzene were located in FEDRIP 2001. 

The Toxic Substances Research Initiative of the Government of Canada is sponsoring a study titled Risk 

assessment for hexachlorobenzene: mechanism of gender related tumor promotion for fiscal years 

1999–2002 by researchers Michel Charbonneau and Daniel Cyr at the Universite du Quebec. 

National Institute of Environmental Science (NIEHS) is sponsoring a study for fiscal years 1996–2002 by 

researcher Gayle Windham at the Public Health Institute in Berkeley, California to investigate whether 

there is a relationship between exposure to chlorinated chemicals, which mediate toxicity through the 

neuroendocrine axis in animals, and menstrual cycle disturbances in women in a Laotian immigrant 

population in the San Francisco Bay area. 

NIEHS is sponsoring a study for fiscal years 1996–2002 by researcher Stephen Schwartz at the Fred 

Hutchinson Cancer Research Center in Seattle, Washington to examine the potential relationship between 

exposure to organochlorine compounds and development of uterine fibroids in a case-control study. 

NIEHS is also sponsoring a study for fiscal years 2001–2005 by researcher Stephen Schwartz to 

investigate the potential relationship between exposure to organochlorine compounds and the 

development of testicular germ cell carcinoma in a case-control study. 

NIEHS is sponsoring a program for fiscal year 2001 by researcher Maria Schymura at the State 

University of New York in Albany to monitor organochlorine exposure in a Native American population 

in the state of New York. 

NIEHS is sponsoring a study for fiscal years 2000–2005 by researcher Lawrence Schell at the State 

University of New York at Albany in Albany, New York, to investigate whether there is a relationship 

among exposure to environmental pollutants, thyroid function, and psychosocial outcomes in a Native 

American population in the state of New York. 
HEXACHLOROBENZENE 187 

4. CHEMICAL AND PHYSICAL INFORMATION
 

4.1 CHEMICAL IDENTITY 

Hexachlorobenzene is a chlorinated hydrocarbon industrial chemical. Although hexachlorobenzene is not 

currently manufactured as a commercial end product in the United States, it is formed as a waste product 

in the production of several chlorinated hydrocarbons such as tetrachloroethylene, trichloroethylene, and 

carbon tetrachloride, and is a contaminant in some pesticides such as pentachloronitrobenzene and 

pentachlorophenol. Its presence in the environment is also due to its previous use as a fungicide. 

Hexachlorobenzene is a very persistent environmental chemical due to its chemical stability and 

resistance to biodegradation. Information regarding the chemical identity of hexachlorobenzene is 

located in Table 4-1. 

4.2 PHYSICAL AND CHEMICAL PROPERTIES 

Hexachlorobenzene is a white, crystalline solid (Verschueren 1996) that is practically insoluble in water 

(Lide 1998). When heated to decomposition, it emits toxic fumes of chlorides (Sax 1984). Dimethyl 

formamide and hexachlorobenzene react violently above 65 EC (NFPA 1986). Information regarding the 

physical and chemical properties of hexachlorobenzene is located in Table 4-2. 
HEXACHLOROBENZENE 188 

4. CHEMICAL AND PHYSICAL INFORMATION 

Table 4-1. Chemical Identity of Hexachlorobenzene 

Characteristics Information Reference 

Chemical name Hexachlorobenzene Budavari 1996 

Synonyms Perchlorobenzene; 

HCB; 

pentachlorophenyl chloride 

Budavari 1996 

Farm Chemicals Handbook 2001 

HSDB 2001 

Trade names AntiCarie; Ceku C. B.; 

No Bunt 

Farm Chemicals Handbook 2001 

Chemical formula C6Cl6 Budavari 1996 

Chemical structure Cl 

Cl Cl 

Cl Cl 

Cl 

Identification numbers:

 CAS Registry

 NIOSH RTECs

 EPA Hazardous Waste

 OHM/TADS

 DOT/UN/NA/IMCO 

Shipping

 HSDB

 NCI 

118-74-1 

DA2975000 

U127 

8100010 

UN2729 

1724 

No data 

Farm Chemicals Handbook 2001 

HSDB 2001 

EPA 1999e (40 CFR 261.33) 

HSDB 2001 

HSDB 2001 

NLM 2001 

CAS = Chemical Abstracts Service; DOT/UN/NA/IMO = Department of Transportation/United Nations/North 

America/International Maritime Dangerous Goods Code; EPA = Environmental Protection Agency; 

HSDB = Hazardous Substances Data Bank; NCI = National Cancer Institute; NIOSH = National Institute for 

Occupational Safety and Health; OHM/TADS = Oil and Hazardous Materials/Technical Assistance Data System; 

RTECS = Registry of Toxic Effects of Chemical Substances 
HEXACHLOROBENZENE 189 

4. CHEMICAL AND PHYSICAL INFORMATION 

Table 4-2. Physical and Chemical Properties of Hexachlorobenzene 

Property Information	 Reference 

Molecular weight 284.78 Budavari 1996 

Color White Verschueren 1996 

Physical state Crystalline solid Verschueren 1996 

Melting point 231 EC Budavari 1996 

Boiling point 325 EC Lide 1998 

323–326 EC Budavari 1996 

Density 

at 23 EC 2.044 Budavari 1996 

Odor No data 

Odor threshold:

 Water No data
 

Air No data
 

Solubility:

 Water at 25 EC 0.006 mg/L Farmer et al. 1976 

Water at 20 EC 0.005815 mg/L Yalkowsky 1992 

0.006 mg/L	 Verschueren 1996 

Organic solvents	 Insoluble in water, slightly soluble in Lide 1998
 

ethanol, soluble in ethyl ether, and
 

very soluble in benzene 

Partition coefficients:

 Log octanol/water 5.73 Hansch et al. 1995 

Log Koc	 6.08 EPA 1981a 

5.22	 Kenaga and Goring 1978 

3.59 

Vapor pressure at 20 EC 1.09x10-5 mmHg Budavari 1996 

Henry's law constant 5.8x10-4 atm-m3

/mol Ten Hulscher et al. 1992 

Hydroxyl radical constant 2.7x10-14 cm3

/molecule-second Brubaker and Hites 1998 

at 25 EC 

Autoignition temperature No data 

Flashpoint 242 EC Budavari 1996 

Flammability limits No data 

Conversion factors 1 ppm = 11.8 mg/m3 Verschueren 1996 

1 mg/m3

 = 0.08 ppm 

Explosive limits No data 

HEXACHLOROBENZENE 191 

5. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

5.1 PRODUCTION 

Hexachlorobenzene is not currently manufactured as a commercial end product in the United States, and 

evidence indicates that it has not been commercially produced since the late 1970s (Beyer 1996; EPA 

1986e). However, limited amounts of hexachlorobenzene are produced for laboratory use. The 

compound can be produced commercially by reacting benzene with excess chlorine in the presence of 

ferric chloride at 150–200 EC. In addition, at least one former producer isolated hexachlorobenzene from 

distillation residues obtained as a by-product in the manufacture of tetrachloroethylene (IARC 1979). 

Hexachlorobenzene is currently produced as a by-product or impurity in the manufacture of several 

chlorinated solvents (e.g., tetrachloroethylene, trichloroethylene, carbon tetrachloride); other chlorinated 

compounds (e.g., vinyl chloride); and several pesticides, including pentachloronitrobenzene (PCNB), 

tetrachloroisophthalonitrile (chlorothalonil), 4-amino-3,5,6-trichloropicolinic acid (picloram), pentachlorophenol (PCP) (only in Europe) (EPA 1986e; Tobin 1985), and dimethyltetrachloroterephthalate 

(DCPA or Dacthal®) (Verschueren 1996); and was also produced as a by-product during the production of 

atrazine, propazine, simazine, and mirex (IARC 1979). Hexachlorobenzene is also released in the 

environment due to ongoing use in developing countries and improper storage or disposal in developed 

countries (Dewailly et al. 1999). 

Currently, hexachlorobenzene is produced for on-site use and processing, as a by-product, or as an 

impurity by the following facilities: Albemarle Corporation of Baton Rouge, Louisiana; AMVAC 

Chemical Corporation of Los Angeles, California; Dow Chemical Co. of Freeport, Texas; Dow Chemical 

Co. of Plaquemine, Louisiana; GB Biosciences Corporation of Houston, Texas; Oxy Vinyls LP of La 

Porte, Louisiana; PPG Industries, Inc. of Lake Charles, Louisiana; Vonroll America Inc. of East 

Liverpool, Ohio; Vulcan Materials Co. of Geismar, Louisiana (TRI99 2001). Hexachlorobenzene is also 

processed by the following waste disposal facilities: Chemical Waste Management of Lake Charles, 

Louisiana; Safety-Kleen of Bridgeport, New Jersey; Safety-Kleen of Deer Park, Texas; Clean Harbors 

Environmental Services of Kimbal, Nebraska; and U.S. Filter Recovery Services of Los Angeles, 

California (TRI99 2001). 

In 1972, hexachlorobenzene produced as a by-product during the production of many other chlorinated 

chemicals was estimated to range from 1,123,500 kg (2,476,868 pounds) to 2,224,900 kg 

(4,905,015 pounds) (IARC 1979). Limited data indicate that hexachlorobenzene was produced at the 
HEXACHLOROBENZENE 192 

5. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

Dover Chemical Company, Dover, Ohio and Hummel Chemical Company, South Plainfield, New Jersey, 

until 1977. It is estimated that 1,450 kg (3,200 pounds) of hexachlorobenzene (as an end-product) were 

produced in the United States in 1975, and that 3,500–11,500 kg (7,700–25,350 pounds) of hexachlorobenzene were inadvertently produced in the manufacture of chlorinated solvents in 1984 (EPA 1986e). 

No current estimates of hexachlorobenzene production are available (SRI 2001). 

Table 5-1 lists the facilities in each state that manufacture or process hexachlorobenzene, the intended 

use, and the range of maximum amounts of hexachlorobenzene that are stored on-site. There are 

20 facilities that produce or process hexachlorobenzene in the United States. Current estimates for the 

amounts of hexachlorobenzene stored on-site as a by-product or impurity range from 0 to 999,999 pounds 

per year (0–453,951 kg/year) (TRI99 2001). The data from the Toxics Release Inventory (TRI) listed in 

Table 5-1 should be used with caution, however, since only certain types of facilities were required to 

report (EPA 1995c). This is not an exhaustive list. 

There are some indications that any process that produces dioxins or dibenzofurans (e.g., pulp and paper 

mills using chlorine for bleaching) will also yield other chlorinated organic compounds such as hexachlorobenzene (EPA 1992b). In addition, hexachlorobenzene may be produced as a by-product in waste 

streams of chlor-alkali plants and wood preserving plants (Leger 1992), and in fly ash (Eicman et al. 

1981) and flue gas effluents from municipal incineration (Oberg and Bergstrom 1985; Oehme et al. 1987; 

Tiernan et al. 1985). No estimates for the amount of hexachlorobenzene produced as a by-product via 

these sources was available. 

5.2 IMPORT/EXPORT 

Hexachlorobenzene imports of 2,440 kg (5,400 pounds) in 1977 and 17,300 kg (38,100 pounds) in 1982 

were reported (HSDB 1995). No other recent data for import or export in the United States was located. 

However, the United Nations Environment Program conducted a survey of import/export data for 

persistent organic pollutants worldwide. Although this report contained no production data, it did contain 

worldwide import/export data. From 1990–1994, 158.45 tons of hexachlorobenzene were exported 

worldwide, while 2,258 tons of hexachlorobenzene were imported worldwide (UNEP 1996). It should be 

noted that U. S. import/export data are not included in this study. 
HEXACHLOROBENZENE 193 

5. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

Table 5-1. Facilities that Produce, Process, or Use Hexachlorobenzene 

Number of Minimum amount on Maximum amount on 

Statea facilities site in poundsb site in poundsb Activities and usesc 

CA 4 100 99,999 1, 2, 3, 5, 6, 13 

KY 2 1,000 99,999 1, 3, 7, 13 

LA 5 0 999,999 1, 3, 5, 7, 8, 13 

NE 1 10,000 99,999 13 

NJ 1 1,000 9,999 13 

OH 1 1,000 9,999 13 

TN 1 1,000 9,999 1, 6 

TX 5 1,000 999,999 1, 3, 5, 6, 13 

Source: TRI99 2001 

a

Post office state abbreviations used 

b

Amounts on site reported by facilities in each state c

Activities/Uses: 

1. Produce 6. Impurity 10. Repackaging 

2. Import 7. Reactant 11. Chemical processing aid 

3. Onsite use/processing 8. Formulation 12. Manufacturing aid 

4. Sale/Distribution component 13. Ancillary/other uses 

5. Byproduct 9. Article component 
HEXACHLOROBENZENE 194 

5. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

5.3 USE 

There are no current commercial uses of hexachlorobenzene as an end-product in the United States. 

However, hexachlorobenzene was used as a fungicide on the seeds of onions, sorghum, wheat, and other 

grains (IARC 1979) until 1984, when the last registered use of the compound as a pesticide was 

voluntarily cancelled. Hexachlorobenzene was also used in the production of pyrotechnic and ordinance 

materials for the military, the production of synthetic rubber (EPA 1986e), as a porosity controller in the 

manufacture of electrodes, a chemical intermediate in dye manufacturing, and a wood preservative (IARC 

1979). 

5.4 DISPOSAL 

Hexachlorobenzene is listed as a toxic substance under Section 313 of the Emergency Planning and 

Community Right-to-Know Act (EPCRA) under Title III of the Superfund Amendments and 

Reauthorization Act (SARA) (EPA 1995c). Disposal of wastes containing hexachlorobenzene is 

controlled by a number of federal regulations. Land disposal restrictions (e.g., treatment standards) apply 

to wastes containing hexachlorobenzene. Incineration at high temperatures is a proposed disposal 

method, but incineration can lead to chlorinated products as toxic as hexachlorobenzene. Past disposal 

methods for industrial wastes containing hexachlorobenzene included incineration, disposal in landfills, 

discharge to municipal sewage treatment plants, and emission to the atmosphere (Clayton and Clayton 

1981; EPA 1988a, 1989a). High temperature incineration (around 1,300 EC) with a retention time of 

approximately 0.25 seconds is the recommended disposal technique because it is reported to destroy more 

than 99% of the chemical (IRPTC 1985). Lamb et al. (1994) reported that organic waste compounds 

including hexachlorobenzene are used to co-fire cement kilns. These authors reported that the destruction 

efficiency of hexachlorobenzene fed to a rotary kiln/afterburner incinerator was greater than 99.9999%. 

Landfill disposal of hexachlorobenzene can lead to migration of the compound via water and sublimation 

of the compound into the air. Calaminus et al. (1993) conducted pyrolysis experiments with hexachlorobenzene in an inert atmosphere of argon. These authors reported hexachlorobenzene was substantially 

pyrolyzed (70%) at temperatures of 1,100 EC for 20 seconds into elemental carbon (soot) and chlorine 

(Cl2), but that other polychlorinated compounds (e.g., hexachlorohexane, hexachloro-1,3-cyclopentadiene, octachlorostyrene, octachloronaphthalene, octochloroacenaphthalene, and decachloronaphthoacenaphthalene) were also produced. Process wastes containing hexachlorobenzene from the 

production of chlorinated aliphatic hydrocarbons has an EPA-prescribed treatment standard before land 
HEXACHLOROBENZENE 195 

5. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

disposal. These wastes must be treated to specified concentrations prior to land disposal at a hazardous 

waste facility (EPA 1995c). Deep well injection of HCB is also not recommended (IRPTC 1985). 

The waste water treatment technology that most closely resembles incineration is wet air oxidation. It is 

specifically designed to destroy organics in waste waters and efficiently oxidizes organics in aqueous 

media by operating at relatively high temperatures and pressures. Furthermore, wet air oxidation is 

typically performed on waste waters that contain relatively high concentrations of organics (i.e., those that 

are at or near the 1% total organic carbon cutoff for waste water). Carbon adsorption has been specified 

as part of the treatment train because hexachlorobenzene is believed to be adsorbable when present in low 

concentrations as might be expected in an effluent from either wet air or chemical oxidation (EPA 1985c). 

No other information was located on the past or present volumes of hexachlorobenzene disposed of by 

each disposal method. 

HEXACHLOROBENZENE 197 

6. POTENTIAL FOR HUMAN EXPOSURE 

6.1 OVERVIEW 

Hexachlorobenzene has been identified in at least 107 of the 1,613 hazardous waste sites that have been 

proposed for inclusion on the EPA National Priorities List (NPL) (HazDat 2002). However, the number 

of sites evaluated for hexachlorobenzene is not known. The frequency of these sites can be seen in 

Figure 6-1. All of these sites are located within the United States and none are located in the 

Commonwealth of Puerto Rico (not shown). 

Hexachlorobenzene has not been sold in the United States (as an end-product) since its last registered use 

as a pesticide was voluntarily canceled in 1984. Hexachlorobenzene is released to the environment 

almost entirely by industrial activities. Currently, it is released as a by-product during the manufacture of 

several chlorinated solvents (carbon tetrachloride, perchloroethylene, trichloroethylene, and chlorinated 

benzenes) or as an impurity in several currently registered pesticides. Additional amounts of hexachlorobenzene are also formed during combustion processes such as incineration of municipal refuse (EPA 

1986e) or through use of pyrotechnic mixtures (Karlsson et al. 1991). 

Hexachlorobenzene may enter the environment through air emissions and in waste water from facilities 

involved in the production of solvents or pesticides, combustion products (i.e., flue gases and fly ash) 

from municipal incinerators, and air emissions released from the use of pyrotechnic mixtures. Nonpoint 

source dispersal of hexachlorobenzene in both agricultural and urban settings results from its presence as 

a contaminant in several widely used pesticides. These sources, in addition to hazardous waste site 

sources, account for the majority of human exposures to hexachlorobenzene. 

Hexachlorobenzene is among the most persistent environmental pollutants because of its chemical 

stability and resistance to degradation. If released to the atmosphere, hexachlorobenzene exists primarily 

in the vapor phase and degradation is extremely slow. Half-life estimations for hexachlorobenzene in the 

atmosphere are highly variable, ranging from 0.63 years in tropical/subtropical regions, to 1.94 years in 

temperate/boreal regions, to 6.28 years in polar regions. A calculated half-life of 1.69 years was attained 

from a measured hydroxyl rate constant (2.7x10-14 cm3

/molecule-second). Long-range global transport is 

possible from the temperate to the polar regions. Physical removal of hexachlorobenzene from the air 
Figure 6-1. Frequency of NPL Sites with Hexachlorobenzene Contamination 

1-2 3-4 5-6 7-8 18 

Frequency of

NPL Sites 

Derived from HazDat 2002
 
HEXACHLOROBENZENE 199 

6. POTENTIAL FOR HUMAN EXPOSURE 

may occur via washout by rainfall or snowfall, or via dry deposition. If released to water, hexachlorobenzene will partition from the water column into sediment and suspended particulate matter. 

In water, it is a persistent chemical not readily degraded by either abiotic or biotic processes. The halflife value of hexachlorobenzene is estimated to range from 2.7 to 5.7 years in surface water and from 

5.3 to 11.4 years in groundwater. Volatilization from the water column is moderately rapid; however, the 

compound's strong adsorption to particulates and organic matter in water can result in lengthy persistence 

in the sediment. If released to soil, hexachlorobenzene can volatilize from the soil surface relatively 

quickly, but will be strongly adsorbed to organic matter and is generally considered immobile with 

respect to leaching. Its half-life value in soils is estimated to range from 3 to 6 years. Hexachlorobenzene 

bioaccumulates significantly in both terrestrial and aquatic food chains. Bioconcentration factors (BCFs) 

as high as 17,000,000 and 21,900 have been reported for lichens and fish, respectively. 

Monitoring for hexachlorobenzene has focused primarily in the Great Lakes region where production of 

chlorobenzenes was historically high. Atmospheric monitoring detected the compound at mean and 

median concentrations of 36.68 (0.03668 ng/m3

) and 30.94 (0.03094 ng/m3) pg/m3, respectively, from 

56 air samples in Villeroy, Quebec in 1992 (Poissant et al. 1997). Hexachlorobenzene has also been 

detected in minute amounts (up to 0.174 ng/L [ppt]) in precipitation samples from the Great Lakes region 

(Chan et al. 1994) and in precipitation samples collected from Villeroy, Quebec in 1992 (0.04 ng/L) 

(Poissant et al. 1997). It was also detected in drinking water in three cities on Lake Ontario at a mean 

concentration of 0.1 ppt (Oliver and Nichol 1982a). Hexachlorobenzene has also been detected in soil 

and sediment samples in both agricultural areas where it was used as a fungicide on seed grains and in 

urban soils near production and waste disposal sites (EPA 1985g). Sediment samples (2 cm depth) 

collected from lakes (Allen-Gil et al. 1997a) and landfills have also been contaminated with hexachlorobenzene (Yasuhara et al. 1999). 

Concentrations of hexachlorobenzene have been reported for a variety of commercial fish species in the 

Great Lakes (Allen-Gil et al. 1997a; Kuchlick and Baker 1998) with concentrations up to 17 ppb in raw 

fish fillets (Newsome and Andrews 1993; Zabik et al. 1995). In the National Pesticide Monitoring 

Program, concentrations as high as 700 ppb were reported in whole fish samples collected from 

contaminated areas (Schmitt et al. 1990). This chemical has also been detected in the fatty tissues and 

muscle of a wide variety of waterfowl (Foley 1992; Gebauer and Weseloh 1993; Swift et al. 1993), 

marine mammals (Becker et al. 1997; Langlois and Langis 1995), and mammals (Corsolini et al. 1999). 

In terrestrial ecosystems, hexachlorobenzene has been detected in lichens (Muir et al. 1993) and in 
HEXACHLOROBENZENE 200 

6. POTENTIAL FOR HUMAN EXPOSURE 

caribou that graze primarily on lichens (Elkin and Bethke 1995). Concentrations of hexachlorobenzene in 

these fish and wild game species can be a source of hexachlorobenzene exposure to man. 

Hexachlorobenzene residues have been detected in 76% of samples analyzed as part of the National 

Human Adipose Tissue Survey (FY82) (EPA 1986f). These hexachlorobenzene residues are most likely 

the result of consumption of low levels of hexachlorobenzene in food, with a calculated yearly intake of 

68, 22, and 5 μg for adults, toddlers, and infants, respectively (EPA 1986e). Compared to this, exposure 

to hexachlorobenzene via inhalation or through drinking water is relatively low. Human exposure may 

also occur via dermal contact with contaminated soil or sediment or via ingestion of contaminated soil by 

children. In occupational settings, exposure occurs primarily via inhalation or dermal contact. 

Hexachlorobenzene has been identified in at least 107 of 1,613 current or former NPL hazardous waste 

sites (HazDat 2002). However, the number of sites evaluated for hexachlorobenzene is not known. The 

frequency of these sites within the United States can be seen in Figure 6-1. 

Due to extensive research conducted on hexachlorobenzene, the data reported herein do not encompass 

complete and thorough research for this chemical. 

6.2 RELEASES TO THE ENVIRONMENT 

Releases of hexachlorobenzene are required to be reported under SARA Section 313; consequently, data 

are available for this compound in the Toxics Release Inventory (TRI) (EPA 1995c). According to the 

TRI, a total of 13,818 pounds (6,267 kg) of hexachlorobenzene was released to the environment in 1998. 

In addition, an estimated 490 pounds (222 kg) were released by manufacturing and processing facilities to 

publicly owned treatment works (POTWs) and an estimated 13,328 pounds (6,045 kg) were transferred 

off-site (TRI98 2000). The TRI data should be used with caution because only certain types of facilities 

are required to report. This is not an exhaustive list. 

Hexachlorobenzene has been identified in a variety of environmental media (air, surface water, leachate, 

groundwater, soil, and sediment) collected at 107 of 1,613 current or former NPL hazardous waste sites 

(HazDat 2002). 
HEXACHLOROBENZENE 201 

6. POTENTIAL FOR HUMAN EXPOSURE 

6.2.1 Air 

Releases to the air from the production of chlorinated solvents, where hexachlorobenzene is a minor byproduct, have been estimated at 0.3 kg (0.7 pounds) annually and emissions from municipal refuse 

incineration have been estimated at 85–8,512 kg (187–1,870 pounds) annually (Bailey 2001). 

Table 6-1 Part B summarizes U.S. hexachlorobenzene emissions in the mid 1990s. As incineration has 

emerged as a prevalent technology for reducing the bulk of hazardous and nonhazardous wastes, 

investigations have shown that even with careful controls it is virtually impossible to eliminate unwanted 

by-products (Products of Incomplete Combustion [PICs]) (Martens et al. 1998). Slight temperature 

differences on the surfaces of incinerator kiln and reactor components, or other reactions in flues, can lead 

to the formation of numerous chemical compounds. Where the original wastes contain organochlorines, 

one type of toxicant may be transformed into another (Dellinger et al. 1989). Hexachlorobenzene, for 

example, has been detected at concentrations ranging from 20 to 70 ppm in gases emitted by the thermal 

degradation of toxaphene (Lahaniatis et al. 1992). Since incineration of wastes is a growing global 

phenomenon, there are concerns that inadequate management attention is given to minimizing PICs. For 

hexachlorobenzene, and many other organochlorines that can be dispersed widely through atmospheric 

transport pathways, the virtual absence of data on PICs can lead to complications in estimating 

environmental releases and mass balances of hexachlorobenzene for regional areas or on a global scale 

(Lahaniatis et al. 1992). 

Nonpoint source dispersal of hexachlorobenzene historically has resulted from its use as a seed fungicide 

(Beyer 1996) and results from the use of a number of registered pesticides in which it is a contaminant. 

Eight major pesticides (chlorothalonil, DCPA or Dacthal®, pentachlorophenol or PCP, picloram, PNCB or 

quintozene, atrazine, simazine, and lindane) in current use contain up to 0.3% hexachlorobenzene as an 

impurity (see Section 6.2.3 and Table 6-1 Part A). When these pesticides are applied in sprays, they have 

the greatest potential for release into the air. Most of the pesticide, and the hexachlorobenzene impurities, 

end up on the top layer of the soil and can become airborne through volatilization of the vapor or 

adsorbed onto soil particles. The hexachlorobenzene agriculturally applied through the use of these eight 

pesticides amounts to an estimated 1,270 kg/year (2,790 pounds/year); however, the total amount of hexachlorobenzene actually released into the air could not be estimated (Bailey 2001). 

Another minor source of hexachlorobenzene releases to the air comes from the use of pyrotechnic 

mixtures that produce white obscurant screening smokes (Karlsson et al. 1991). These screening smokes 
Table 6-1. U.S. Hexachlorobenzene Emissions Summary, Mid 1990s, Part A 

HCB Average annual HCB emissio

concentration U.S. use (kg/year) (kg/year) 

Products containing HCB Pesticides 

DCPA (dacthal) 1,000 ppm 677,791 

Atrazine 1 ppm 32,233,523 

Simazine 1 ppm 1,825,391 

Picloram 50 ppm 655,922 

PCNB 500 ppm 627,338 Chlorothalonil 40 ppm 4,928,591 

Lindane 50 ppm 51,345 

Pentachlorophenol 50 ppm 4,000,000 

Pesticides subtotal 

Chlorinated solvent usage 1 ppb 282,600,000 

Manufacturing 

U.S. Chemical manufacturing 
Table 6-1. U.S. Hexachlorobenzene Emissions Summary, Mid 1990s, Part B 

Emissions factors (kg HCB/kg) Calculated emissi

Product Using Using 

Low Mean High (kg/year) low factors mean fact

Metals 

Aluminum 86 2.2x10-3 4.3x10-3 HCE 5.88x104 51 12

casting Secondary 3.9x10-9 3.9x10-8 3.9x10-7 6.95x108 3 2

copper 

Metals subtotals 53 15

Combustion 

Municipal 2.9x10-9 2.9x10-8 2.9x10-7 2.93x109 85 85

Hazardous 1.9x10-9 1.9x10-8 1.9x10-7 1.5x106 3 2

wastes Medical 2.9x10-9 2.9x10-9 2.9X10-9 7.7x108 2 2Coal 1.3x10-12 1.3x10-11 1.3x10-10 7.91x1011 1 Cement 1.7x10-11 1.7x10-10 1.7x10-9 6.76x1010 1 Iron sintering 1.5x10-10 1.5x10-9 1.5x10-8 1.22x1010 2 Sewage sludge 4.7x10-10 4.7x10-9 4.7x10-8 8.65x108 0 

Biomass 6.0x10-12 6.0x10-11 6.0x10-10 2.28x1011 1 

Combustion subtotals 96 96

Total U.S. HCB emissions (kg/year) 1,818 2,78

aBased on 6% volatilization of HCB from pentachlorophenol-treated wood. 

Source: Bailey 2001 
HEXACHLOROBENZENE 204 

6. POTENTIAL FOR HUMAN EXPOSURE 

are used by the military to obscure vision and hide targets, and are used by civilian firefighters during fire 

training sessions. 

The estimated release of 1,535 pounds (698 kg) of hexachlorobenzene to the atmosphere from 

12 manufacturing, processing, and waste disposal facilities in 1999 accounted for about 10.5% of the 

estimated total environmental releases (TRI99 2001). These releases are summarized in Table 6-2. The 

data from the TRI listed in Table 6-2 should be used with caution, however, since only certain types of 

facilities are required to report (EPA 1995c). This is not a comprehensive list. 

Hexachlorobenzene has been identified in air samples at 7 sites, collected from 1,613 NPL hazardous 

waste sites, where it was detected in some environmental media (HazDat 2002). 

6.2.2 Water 

The principal release of hexachlorobenzene into water in the past has been through direct discharges from 

chemical solvent manufacturing facilities. Total production of chlorobenzenes in the United States has 

declined from more than 300,000 metric tons (300 million kg or 661 million pounds) in 1970 to about 

200,000 metric tons (200 million kg or 441 million pounds) in 1980. The total amount of hexachlorobenzene released as a by-product in production of all chlorinated solvents has been estimated to range 

from 70,343 to 241,311 kg/year (154,000–532,000 pounds/year) (EPA 1986e). Estimated hexachlorobenzene releases into water from these sources, however, were only 70 kg/year (154 pounds/year) (EPA 

1986e). 

The estimated release of 8 pounds (4 kg) of hexachlorobenzene to water from two domestic 

manufacturing and processing facilities in 1999 accounted for about 0.055% of the estimated total 

environmental releases (TRI99 2001). An additional 13,559 pounds (6,045 kg) were transferred off-site, 

including to POTWs (TRI99 2001). These releases are summarized in Table 6-2. The data from the TRI 

listed in Table 6-2 should be used with caution, however, since only certain types of facilities are required 

to report (EPA 1995c). This is not a comprehensive list. 

Hexachlorobenzene has been identified in surface and groundwater samples at 10 and 42 sites, 

respectively, collected from 1,613 NPL hazardous waste sites, where it was detected in some 

environmental media (HazDat 2002). 
Table 6-2. Releases to the Environment From Facilities That Produce, Process, or Use Hexa

Reported amounts released in pounds per yeara 

Number 

Stateb 

of facilities Airc Water 

Underground injection Land 

Total on-site released 

Total off-site releasee 

CA 4 7 0 No data 0 7 No data 

KY 2 0 No data No data No data 0 No data LA 5 1,387 7 No data 13,003 14,397 1 

NE 1 5 No data No data 0 5 No data 

NJ 1 3 0 No data 0 3 4 OH 1 2 1 No data No data 3 5 

TN 1 0 No data No data No data 0 No data 

TX 5 131 0 No data 20 151 13,549 

Total 21 1,535 8 0 13,023 14,566 13,559 

Source: TRI99 2001 aData in TRI are maximum amounts released by each facility.bPost office state abbreviations are used.
 cThe sum of fugitive and stack releases are included in releases to air by a given facility. dThe sum of all releases of the chemical to air, land, water, and underground injection wells.
 eTotal amount of chemical transferred off-site, including to publicly owned treatment works (POTW).
 
HEXACHLOROBENZENE 206 

6. POTENTIAL FOR HUMAN EXPOSURE 

6.2.3 Soil 

Historically, hexachlorobenzene was released to soils directly through its application as a fungicide on 

seed grains. Even after use of hexachlorobenzene as a pesticide ceased, an estimated 95% of hexachlorobenzene produced in the manufacture of chemical solvents was disposed of in land applications (EPA 

1986e). Current releases to soils may occur through land disposal of hexachlorobenzene-containing 

wastes, from discharges from manufactured facilities, and from use of currently registered pesticides 

containing small amounts of hexachlorobenzene. Contamination of soil and sediment with hexachlorobenzene is frequently secondary to the discharge of contaminated water, from which the hexachlorobenzene is then adsorbed by the soil or sediment. Sediment samples (2 cm depth) were collected from 

four lakes (Feniak, Desperation, Schrader, and Elusive) from the foothills of the Brooks Range, Alaska. 

All lakes were oligotrophic, and the pH was neutral to slightly alkaline. The mean concentration among 

the six samples was 0.17 ng/g dry weight, derived from the concentrations of the sediments from Lakes 

Feniak (0.27 ng/g dry weight), Desperation (0.08 ng/g dry weight), Schrader (0.11 ng/g dry weight), and 

Elusive (0.21 ng/g dry weight) (Allen-Gil et al. 1997a). Substantial levels of hexachlorobenzene, ranging 

from 0.5 to 460 ppb, have been detected in sediment cores sampled at 1 cm intervals to a sediment depth 

of 8 cm in Lake Ontario. The highest sediment contamination in Lake Ontario was found in sediments 

1–2 cm in depth which correspond to sediments laid down from 1971 to 1976 during a period of high 

U.S. production of chlorobenzenes (Oliver and Nichol 1982a). Although no studies concerning the 

release of hexachlorobenzene from landfills have been located for the United States, Yasuhara et al. 

(1999) sampled leachates from 11 landfills in Japan. Leachates were sampled at the outlet of the leachate 

collecting pipe from open and controlled landfills. Hexachlorobenzene concentrations in the 11 leachate 

samples from 11 landfills ranged from not detected to 0.054 ng/L. Of the 11 sites, 6 are currently under 

reclamation and at least 3 sites were sampled 12–17 years after completion of reclamation. Site 2, which 

was sampled after 14 years of reclamation, had no detectable levels of hexachlorobenzene. Leachates 

from sites 3 and 4, which were sampled after 17 and 12 years of reclamation, had hexachlorobenzene 

concentrations of 0.033 and 0.054 ng/L, respectively. Site 4 had the highest detection of hexachlorobenzene concentration. The median concentration among these sites was 0.03 ng/L. 

The presence of hexachlorobenzene as an impurity in several currently registered pesticides appears to be 

a continuing source of exposure for the general population. The pesticides containing impurities of hexachlorobenzene include: picloram, PCNB or quintozene, chlorothalonil, DCPA or Dacthal®, pentachlorophenol or PCP, atrazine, simazine, and lindane (Bailey 2001; Kutz et al. 1991). Estimated emissions of 

these eight pesticides is summarized in Table 6-1 Part A. Picloram is a herbicide used in agriculture and 
HEXACHLOROBENZENE 207 

6. POTENTIAL FOR HUMAN EXPOSURE 

silviculture to control broad-leaf weeds and conifers in grasses (Tomlin 1997). PCNB is used as a 

fungicide, herbicide, disinfectant, and antifouling ingredient in paint and wood preservatives. It is also a 

seed dressing agent widely used on turf and ornamental plants (EPA 1986e). PCNB, or quintozene, is 

used as a soil fungicide on lawns and ornamental crops, as a seed treatment of field crops and vegetables 

(e.g., barley, corn, cotton, oats, rice, and wheat), and as a slime inhibitor in industrial waters (Budavari 

1996; IARC 1974). Chlorothalonil, and Dacthal® are used widely in agriculture, but also are used in 

home gardens, lawn care, or other applications around residences or in urban areas (Farm Chemicals 

Handbook 1993). Chlorothalonil (sold under the trade name Bravo®) is a fungicide used on horticultural 

crops, golf courses, and residential turf, and as a biocide in paints and wood preservatives. Dacthal® is a 

pre-emergent herbicide widely used on lawns and turf grass (EPA 1986e). Pentachlorophenol (PCP) is a 

insecticide and fungicide used to protect timber from fungal rot and wood-boring insects (Tomlin 1997). 

Atrazine and simazine are selective herbicides used to control broadleaf and grassy weeds in corn and 

other crops (Tomlin 1997). Lindane is an insecticide and fumigant that has been used on a wide range of 

soil-dwelling and plant-eating insects. It is commonly used on a wide variety of crops, in warehouses, in 

public health to control insect-borne diseases, and (with fungicides) as a seed treatment (Tomlin 1997). 

Although hexachlorobenzene impurities in Dacthal® were as much as 10% in the early 1970s, current 

levels of hexachlorobenzene contamination in all five pesticides are much lower. A registration standard 

was issued for picloram in 1985 that specified a maximum hexachlorobenzene content of 0.02%. By the 

terms of an EPA PCNB rebuttal presumption against registration in 1982, PCNB registrants agreed to 

reduce hexachlorobenzene contamination levels in PCNB to 0.5% by 1983 and to 0.1% by April 1988. A 

registration standard was issued by EPA for chlorothalonil in 1984 requiring that hexachlorobenzene 

contamination not exceed 0.05%. Since 1973, the maximum allowable hexachlorobenzene content of 

technical grade DCPA (Dacthal®) has been 0.3% (EPA 1986e). As a result of a settlement agreement 

between the EPA and chemical producers of PCP, the producers agreed to reduce hexachlorobenzene 

contamination in PCP to no more than 75 ppm (0.0075%). Recent surveillance monitoring by EPA has 

generally detected <50 ppm (0.005%) hexachlorobenzene contamination in PCP samples (EPA 1994c). 

In 1999, 13,023 pounds of hexachlorobenzene were released to land from 21 domestic manufacturing, 

processing, and waste disposal facilities reporting releases of the compound to the environment (TRI99 

2001). No releases (0 pounds) of hexachlorobenzene occurred via underground injection (TRI99 2001). 

Releases to the environment from facilities that produce, process, or use hexachlorobenzene are 

summarized in Table 6-2. The data from the TRI should be used with caution since only certain types of 

facilities are required to report (EPA 1995c). This is not a comprehensive list. 
HEXACHLOROBENZENE 208 

6. POTENTIAL FOR HUMAN EXPOSURE 

The annual total emissions of hexachlorobenzene in the United States was estimated to be on average 

2,785 kg (Bailey 2001). This estimated total was calculated using emission subtotals of 1,270 kg for 

pesticide use, 0.3 kg for chlorinated solvent use, 399 kg from manufacturing processes, 156 kg from metal 

industries (using mean emission factors), and 960 kg from combustion sources (using mean emission 

factors). These data are summarized in Table 6-1. 

Hexachlorobenzene has been identified in soil and sediment samples at 76 and 33 sites, respectively, 

collected from 1,613 NPL hazardous waste sites, where it was detected in some environmental media 

(HazDat 2002). 

6.3 ENVIRONMENTAL FATE 

6.3.1 Transport and Partitioning 

Hexachlorobenzene has a moderate vapor pressure (1.09x10-5 mmHg) (Budavari 1996) and has a very 

low solubility (0.005815 mg/L) (Yalkowsky 1992) in water (see Table 4-2). If released to the 

atmosphere, hexachlorobenzene can exist in both the vapor phase in association with particulates 

(Eisenreich et al. 1981); however, monitoring studies have demonstrated that the vapor phase generally 

predominates (Ballschmiter and Wittlinger 1991; Bidleman et al. 1989; Lane et al. 1992). Hexachlorobenzene concentrations in the vapor phase represented 92–100% of the total hexachlorobenzene 

concentration in air samples collected in a monitoring study conducted in Ontario, Canada (Lane et al. 

1992). Although physical removal of hexachlorobenzene from the atmosphere is possible via both wet 

and dry deposition (Howard 1990), the compound is hydrophobic, and somewhat resistant to wet 

deposition scouring unless it becomes sorbed to airborne dust or cloud condensation nuclei. Its resistance 

to wet deposition tends to slow down its transfer across the equatorial areas between the northern and 

southern hemispheres. At high latitudes, the typically cold air conditions encourage dry deposition of 

aerosols (Ballschmiter and Wittlinger 1991; Lane et al. 1992; Wania and Mackay 1993). These factors 

lead to atmospheric pathways as a major transport mechanism for hexachlorobenzene. The net residence 

time of hexachlorobenzene in the atmosphere is significantly less than 1 year, and is based on physical 

translocation and not on chemical transformation (Ballschmiter and Wittlinger 1991). The atmospheric 

mechanisms can operate over large distances, perhaps on a hemispheric scale (Kelly et al. 1991). At 

higher latitudes, transfers and partitioning back to aqueous systems may be accentuated by the cold, dry 

air. Since these areas are not zones of hexachlorobenzene production or use, the presence of such 

contaminants has attracted considerable attention in research studies (Ballschmiter and Wittlinger 1991). 
HEXACHLOROBENZENE 209 

6. POTENTIAL FOR HUMAN EXPOSURE 

The Henry's law constant value for this compound (5.8x10-4 atm-m3/mol) (Ten Hulscher et al. 1992) 

suggests that releases of hexachlorobenzene to surface water will volatilize at a moderate rate, and that 

volatilization can be a significant transfer mechanism (Thomas 1990). If released to water, adsorption of 

hexachlorobenzene to sediment or soil particulates is also expected to be significant on the basis of the 

high organic carbon partition coefficient (Koc) value (EPA 1981a) (see Table 4-2). Since hexachlorobenzene will adsorb strongly to soil particles and sediments, it may build up in the bottom sediments of 

large aquatic systems such as the Great Lakes. The concentration of hexachlorobenzene in Lake Ontario 

sediment is about 1 million times higher than in Lake Ontario water (Oliver and Nichols 1982a). In 

Lakes Superior, Michigan, and Huron, the very large sizes, considerable depths, and appreciable retention 

times have allowed much of the historical organochlorine burden to become immobilized in bottom 

sediments, with a concomitant reduction in the levels of hexachlorobenzene found in surface waters. In 

the upper Great Lakes, the vast majority of the ongoing mass balance inputs seem related to atmospheric 

deposition (Eisenreich et al. 1981). For other parts of the Great Lakes system, and especially the 

connecting waters and Lakes Erie and Ontario, mass balance studies can give different impressions. 

Detailed studies on Lake Erie suggest that well over half of the hexachlorobenzene inputs come from wet 

or dry atmospheric deposition processes (270 kg/year or 600 pounds/year) (Kelly et al. 1991). However, 

a significant portion (110 kg/year or 240 pounds/year) also comes from river pathways such as the highly 

polluted Detroit River via surface runoff and contaminated sediments resuspended during their passage 

through the connecting waters. Because of its strong adsorption to sediment, most of the hexachlorobenzene is transported with silt and sediment particles during floods, and very little is dissolved in the 

water. Quemerais et al. (1994) reported that 23% of the hexachlorobenzene in whole water samples 

collected from the St. Lawrence River was associated with the dissolved phase, while 77% was associated 

with the particulate phase. Although Rostad et al. (1988) did not quantify the percentage of hexachlorobenzene found in the dissolved versus the particulate phase, they reported that hexachlorobenzene was 

one of the organic compounds associated with suspended sediment particles in several river systems 

within the Mississippi River drainage area. In a 1999 study, Rostad et al. (1999) measured hexachlorobenzene concentrations in suspended sediment particles within the Mississippi River in the spring and 

summer of 1989 and 1990. Concentrations of hexachlorobenzene fluctuated between Winfield, Missouri 

and Belle Chasse, Louisiana during the spring and the summer; however, in both seasons, the 

concentration was higher at Chasse, Louisiana (1.9 ng/g in the spring; 2.1 ng/g in the summer) than at 

Winfield, Missouri (1.2 ng/g in the spring; 0.13 ng/g in the summer). Furthermore, Rostad et al. (1999) 

estimated annual transport of hexachlorobenzene via suspended sediments to the Gulf of Mexico by 

averaging the St. Francisville, Louisiana and Belle Chasse, Louisiana daily loads, averaging the spring 

and summer values for hexachlorobenzene, and projecting annual transport. The Gulf of Mexico receives 
HEXACHLOROBENZENE 210 

6. POTENTIAL FOR HUMAN EXPOSURE 

an estimated 145 kg/year hexachlorobenzene via suspended sediment particles of the Mississippi River 

(Rostad et al. 1999). 

Because of its high sorption characteristics, hexachlorobenzene is expected to be immobile in soil and 

unlikely to leach into groundwater (Swann et al. 1983). At waste disposal sites, where bioremediation 

techniques are proposed to reduce the mass of carbon-containing contaminants, there is the potential for 

augmenting the leaching properties of hexachlorobenzene and other organochlorines. The lipid materials 

in bacterial cell membranes may lead to a repartitioning of hexachlorobenzene sorbed to soil colloids. 

This can lead to a phenomenon called facilitated transport where the mobility of hydrophobic pollutants 

such as hexachlorobenzene adsorbed to soils may be enhanced by biosorption on bacteria and move into 

aquifers along with the bioremedial bacterial cultures (Lindqvist and Enfield 1992). Except at NPL sites, 

however, this potential source of groundwater pollution would seem to be remote. 

The Henry's law constant value (5.8x10-4 atm-m3

/mol) (Ten Hulscher et al. 1992) suggests that hexachlorobenzene released to moist soil will volatilize at a moderate rate. Several studies have indicated that 

volatilization may be a significant mechanism for loss of hexachlorobenzene released to soils. Beall 

(1976) studied the persistence of aerially applied hexachlorobenzene (equivalent to 10 ppm in the top 

5 cm of soil) in a simulated pasture maintained for 19 months in a greenhouse. Twenty hours postapplication, the top 2 cm of soil contained 5.6 ppm (air-dry basis). Hexachlorobenzene concentrations in 

the top 2 cm of soil found after 0.5, 1, 6.5, 12, and 19 months were 45.2, 24.4, 7.9, 4.7, and 3.4% of 

day 1 values, respectively. However, no significant change in the deeper 2–4 cm layer of soil which 

averaged hexachlorobenzene residues of 0.11 ppm was observed over the 19-month study. 

Concentrations of hexachlorobenzene in pasture grass on day 1 were 1,060 ppm, but 2 weeks 

postapplication only 15.6 ppm (1.5% of day 1 residues) were detected. Although hexachlorobenzene 

volatilized fairly rapidly from plant and soil surfaces, it could be persistent within the soil if treated 

surface soil were mixed into deeper soil layers by plowing. Nash and Gish (1989) studied the 

volatilization and dissipation of several halogenated pesticides from moist sandy loam soil under 

controlled conditions in micro-agroecosystem chambers maintained in a greenhouse for 154 days. As soil 

temperature increased from 5 to 35 EC, the percentage of originally applied hexachlorobenzene that was 

detected in the soil compartment decreased, while the percentage detected in the air increased suggesting 

that hexachlorobenzene volatilizes more rapidly with increased soil temperature. 

The high octanol/water partition coefficient (Kow) value (Hansch et al. 1995) for hexachlorobenzene (see 

Table 4-2) suggests that bioconcentration and biomagnification of hexachlorobenzene are likely to occur 
HEXACHLOROBENZENE 211 

6. POTENTIAL FOR HUMAN EXPOSURE 

to a significant degree. Veith et al. (1979) measured biological concentration factor (BCF) values of 

16,200 for fathead minnows, 21,900 for green sunfish, and 5,500 for rainbow trout exposed to hexachlorobenzene at 15 EC for 32 days. Oliver and Niimi (1983) studied bioconcentration in rainbow trout 

exposed to water containing 2 concentrations of hexachlorobenzene (0.32 and 8 ng/g [ppb]) for 119 and 

105 days, respectively. The BCF values were 12,000 and 20,000 at the 0.32 and 8 ng/g (ppb) exposure 

levels, respectively. Chaisuksant et al. (1997) conducted a bioconcentration experiment using mosquito 

fish (Gambusia affinis) as well. The fish were exposed to three concentrations of eight chemicals, and the 

highest concentration used consisted of a mixture with each chemical present in a concentration equal to 

1/20 of the LC50. After 96 hours of exposure, the BCF of hexachlorobenzene in mosquito fish was 3,730. 

In 1992, the Chemicals Inspection and Testing Institute (1992) conducted a bioconcentration experiment 

using carp (Cyprinus carpio). After an 8-week exposure period to concentrations of 0.5 and 0.05 μg/L of 

hexachlorobenzene, the BCFs in carp were 11,000–27,000 and 6,000–30,000, respectively. 

In a model aquatic ecosystem to which hexachlorobenzene was introduced, the BCF averaged 740 for 

algae (Oedogonium cardiacum), 1,500 for the snail (Helisoma sp.), 910 for the daphnid (Daphnia 

magna), 1,610 for the mosquitofish (G. affinis), and 10,610 for the catfish (Ictalurus punctatus) (Isensee 

et al. 1976). The authors concluded that biomagnification was also occurring within the food chain 

because the catfish (highest trophic level species) accumulated over 10 times more hexachlorobenzene 

than the next highest trophic level (snails and mosquitofish), and these species accumulated 1.5–2 times 

more than the lowest food chain species, the daphnids (primary consumers) and the algae (primary 

producers). In studies of natural populations of white bass in Lake Erie, Russell et al. (1995) concluded 

that biomagnification of hexachlorobenzene did not occur. These authors did report biomagnification in 

Lake Erie fish populations was occurring for several other organic chemicals with log Kow values greater 

than 6.1. Hexachlorobenzene bioaccumulation factors (BAFs) in aquatic fish species has been measured 

by Burkhard et al. (1997) in the Bayou d’Inde of the Calcasieu River system near Lake Charles, 

Louisiana. This field study resulted in log BAF values of 5.80 for blue crab (Callinectes Sapidus), 

6.03 for mummichog fish (Fundulus Heteroclitus), 6.30 for Atlantic croaker (Micropoganias undulatus), 

and 6.68 for gulf menhaden (Brevoortia Patronus). The author further compared the measured values 

obtained to previously reported and predicted BAF values. A comparison of these data with that of 

Pereria et al. (1988) reveals a difference that was not considered significant by the author. Pereria et al. 

(1988) determined log BAF values of 4.03 for blue crab, 4.56 for Atlantic croaker, 4.12 for spotted sea 

trout (Cynoscion nebulosis), and 4.61 for blue catfish (Ictalurus furcatus). 
HEXACHLOROBENZENE 212 

6. POTENTIAL FOR HUMAN EXPOSURE 

Connell et al. (1988), using data derived from terrestrial laboratory microcosm studies with two 

oligochaete worms (Limnodrilus hoffmeisteri and Tubifex), suggest that interstitial water may be the 

source from which lipophilic compounds such as hexachlorobenzene in sediment are bioconcentrated by 

oligochaetes. The concentration factor was 0.54 for hexachlorobenzene during a 110-day exposure test. 

In a similar study of the earthworm (Eisenia andrei) raised in field-contaminated soil, Belfroid (1995) 

reported a biota-to-soil accumulation factor of 0.507 for hexachlorobenzene. These authors also noted an 

initial elimination half-life of 1.9 days followed by a period of slower elimination with a half-life of 

47 days. In a terrestrial food web study conducted on the Niagara Peninsula of Ontario, Canada from 

1987 to 1989, Hebert et al. (1994) reported concentrations of hexachlorobenzene increased from the 

lower trophic level species to higher trophic level predator species. Concentrations of hexachlorobenzene 

were not detected in soil or plant material; however, concentration ranges were 0.2–0.3 μg/kg (ppb) (wet 

weight) in earthworms, not detected to 1.0 μg/kg (ppb) in mammals, 2.0–2.4 μg/kg (ppb) in starlings, 

1.8–2.5 μg/kg (ppb) in robins, and 2.1–5.1 μg/kg (ppb) in kestrels at the top of the food web. 

Several agricultural species of plants have been shown to bioaccumulate hexachlorobenzene in their roots 

and in portions of the plant growing closest to the soil (Scheunert et al. 1983; Smelt and Leistra 1974). 

There were marked differences in the BCFs among the various plant species with higher residues 

associated with those species with the higher lipid content (Schroll et al. 1994; Smelt and Leistra 1974). 

The roots of the plants generally accumulate higher concentrations of soil-applied hexachlorobenzene 

than do the aerial parts of the plants. This has been demonstrated for hexachlorobenzene in sugar beets, 

carrots, turnips, wheat, and pasture grass (Scheunert et al. 1983; Smelt and Leistra 1974). The edible root 

portion of carrots accumulated the highest concentration of hexachlorobenzene (1,250 ppb with a 

plant/soil BCF of 19) for a human food source. The measured BCF for hexachlorobenzene was 210 and 

470 in soy bean plants via root uptake from water containing 0.2 and 0.4 μg/L hexachlorobenzene, 

respectively (Kraaij and Connell 1997). Concentrations of hexachlorobenzene were also high in grass 

roots (810 ppb) and the lower (0–5 cm) part of the blade (220 ppb) (Smelt and Leistra 1974). It is 

assumed that hexachlorobenzene in soil is mobile mainly in the gas phase. Gaseous hexachlorobenzene 

can diffuse directly into the plant root or evaporated hexachlorobenzene can be taken up by plant foliage 

(Ecker and Horak 1994). Some studies have reported no marked translocation of the hexachlorobenzene 

from roots to shoots or vice versa (Schroll et al. 1994). Residues in the roots were associated only with 

root uptake from the soil; those residues in the shoots were only from foliar uptake from the air. Recent 

studies by Ecker and Horak (1994), however, suggest that root uptake of hexachlorobenzene by oil 

pumpkins occurred and that the hexachlorobenzene was translocated into the shoots. These authors 

believe that uptake of dissolved hexachlorobenzene from soil solution into the roots may not have been 
HEXACHLOROBENZENE 213 

6. POTENTIAL FOR HUMAN EXPOSURE 

considered earlier as a source for the translocated compound. Pollutants entering the plant from 

contaminated soil via roots would be translocated in the plant by the xylem while gas- and particle-phase 

deposition onto leaves or uptake by the stomata would be translocated by the phloem (Simonich and Hites 

1995). Concentrations of hexachlorobenzene in agricultural crops can be directly transferred to humans 

via direct consumption, while concentrations in grass and other forage crops can be indirectly transferred 

to humans via consumption of dairy products or meat from cattle grazing on contaminated pastures. 

Lichen from Northwestern Ontario and South Central Ontario exhibited BCF values of 1.7x107

 and 

8.8x106

, respectively, for hexachlorobenzene. These BCF values were calculated as the concentration of 

hexachlorobenzene in the lichen (ng/m3

 wet weight) compared to the concentration in the air (ng/m3) 

(Muir et al. 1993). Furthermore, bioconcentration of hexachlorobenzene by lichen, a major forage food 

for caribou can transfer hexachlorobenzene to recreational hunters and natives peoples that consume 

caribou in their diets (Elkin and Bethke 1995). 

6.3.2 Transformation and Degradation 

6.3.2.1 Air 

Few studies regarding atmospheric degradation of hexachlorobenzene have been located. Photodegradation of hexachlorobenzene in its vapor phase, or as an adsorbable on silica gel, has been reported as not 

occurring when hexachlorobenzene was irradiated with ultraviolet radiation of wavelength 290 nm for 

6 days (Parlar 1978); however, production of HCl and CO2 was observed when hexachlorobenzene was 

irradiated at 230 nm (Parlar 1978). In the troposphere, hexachlorobenzene is probably photochemically 

stable, but degradation in the stratosphere by photo-dissociation by shorter-wavelength, higher energyultraviolet light may be a mechanism for atmospheric degradation in the stratosphere. 

The photo-oxidation half-life (first-order kinetics) of hexachlorobenzene based on the vapor phase 

reaction with hydroxyl radicals in air was estimated to range from 156.4 days to 4.2 years by Howard et 

al. (1991) and from 158 days to 4.3 years by Kwok and Atkinson (1995). Wania and Mackay (1995) 

estimated the degradation half-life (first-order kinetics) of hexachlorobenzene to be 0.63 years (230 days), 

1.94 years (708 days), and 6.28 years (2,292 days) in air in tropical/subtropical regions, temperate/boreal 

regions, and polar regions, respectively. Brubaker and Hites (1998) measured a hydroxyl rate constant of 

2.7x10-14 cm3

/molecule-second at 25 EC, corresponding to a calculated half-life of 1.69 years. Thus, 

atmospheric degradation is extremely slow and is not an efficient method of hexachlorobenzene removal. 
HEXACHLOROBENZENE 214 

6. POTENTIAL FOR HUMAN EXPOSURE 

6.3.2.2 Water 

Hexachlorobenzene is a persistent compound and is not significantly degraded by either abiotic or 

biodegradation processes in water. It is resistant to the types of hydrolysis reactions that can help degrade 

other organochlorines or organophosphates, and it is not markedly subject to photolytic decay (Mill and 

Haag 1986). Biodegradation of organic priority pollutants in a waste water inoculum system was studied 

by Tabak et al. (1981). Among the 57 environmental pollutants tested, hexachlorobenzene at 

concentrations of 5 and 10 ppm was among the more slowly biodegraded compounds tested, with only 

21–56% degradation in 1 week. Further biodegradation after the first week was minimal in three 

subsequent 7-day subcultures with settled domestic waste water as microbial inoculum. Biodegradation 

of hexachlorobenzene in waste water treatment systems is expected to be slow. 

An aquatic ecosystem study conducted by Schauerte et al. (1982) shows that hexachlorobenzene will 

mainly absorb onto particulate matter in the water and then be transported to the bottom sediment. After 

145 weeks, the study found a significant amount (10–20%) of hexachlorobenzene remaining in the upper 

sediment layers (0–10 cm). The half-life (first-order kinetics) of hexachlorobenzene was estimated to 

range from 2.7 to 5.7 years in surface water and from 5.3 to 11.4 years in groundwater based on 

unacclimated aqueous aerobic biodegradation (Howard et al. 1991). 

Hirsch and Hutzinger (1989) conducted surface water photolysis test with hexachlorobenzene in a 

laboratory setting and found that this process may occur. A first order rate constant (1.3x10-6/sec) 

corresponding to a half-life of 6.17 days for the photolysis of hexachlorobenzene in distilled water in a 

photochemical reactor equipped with mercury arc lamps was reported. Hexachlorobenzene in an 

acetonitrile:water mixture exposed to wavelengths of 290 nm for 8 hours resulted in a 33.5% loss of hexachlorobenzene. 1,2,3,4,5-Pentachlorobenzene (76.8%), 1,2,3,5-tetrachlorobenzene (1.2%), 

1,2,4,5-tetrachlorobenzene (1.7%), and 1,2,4-trichlorobenzene (0.2%) were found as hexachlorobenzene 

transformation products (Choudhry et al. 1986). In another experiment, hexachlorobenzene in a 

water:acetonitrile solution was exposed to sunlight and resulted in a half-life of 70 days (Mill and Haag 

1986). The studies above found photolysis of hexachlorobenzene a feasible loss process with half-lives 

ranging from 6.17 to 70 days. 

Hydrolysis is not expected to be an important fate process. EPA (1987g) observed zero hydrolysis after 

13 days for pH values of 3, 7, and 11 at 85 EC. 
HEXACHLOROBENZENE 215 

6. POTENTIAL FOR HUMAN EXPOSURE 

Hexachlorobenzene can also be eliminated by ozone reactions. Roche and Prados (1995) conducted a 

study and compared the efficiencies of ozone and ozone-hydrogen peroxide systems in removing hexachlorobenzene from water treatment processes. The concentration of ozone during the experiments was 

70 mg O3/L. When ozone was applied, 11–14% of an initial concentration of 1.0 μg/L hexachlorobenzene was removed. This removal increased to 15–48% when hydrogen peroxide and ozone were 

applied together. 

6.3.2.3 Sediment and Soil 

Hexachlorobenzene is a persistent compound and is not significantly degraded in soils by either abiotic or 

biodegradation processes. In a year-long laboratory study, soil treated with 0.1, 1.0, and 10 ppm of hexachlorobenzene was stored under aerobic (sterile and nonsterile) conditions and under anaerobic nonsterile 

conditions in covered containers to retard hexachlorobenzene volatilization (Isensee et al. 1976). No loss 

in the soil-incorporated hexachlorobenzene was observed at any treatment concentration or under any 

storage condition. Beck and Hansen (1974) measured a half-life (first-order kinetics) of 3–6 years for 

hexachlorobenzene in soils. Anaerobic biological dechlorination of hexachlorobenzene has also been 

demonstrated in anaerobic sewage sludge (Fathepure et al. 1988). These authors reported that hexachlorobenzene was dechlorinated to tri- and dichlorobenzenes under anaerobic conditions when sewage 

sludge was maintained in serum bottles and incubated in the laboratory. Complete biotransformation of a 

50 ppm inoculum occurred within 3 weeks. Two routes of dechlorination were observed. The major 

route was hexachlorobenzene 6 pentachlorobenzene 6 1,2,3,5-tetrachlorobenzene 6 1,3,5-trichlorobenzene; the minor route was hexachlorobenzene 6 pentachlorobenzene 6 1,2,4,5-tetrachlorobenzene 6 1,2,4-trichlorobenzene 6 dichlorobenzenes (i.e., 1,2-, 1,3-, and 1,4-dichlorobenzene). Yuan et 

al. (1999) also conducted an anaerobic biological dechlorination study using sewage sludge obtained from 

the Di-Hua Municipal Sewage Treatment Plant in Taipei, Japan. All experiments were performed using 

25 mL serum bottles containing 9 mL of sewage sludge and various concentrations of hexachlorobenzene. 

After a 20-day incubation period, 98% of the 2 mg/L hexachlorobenzene remained, while addition of 

1,2,3-trichlorobenzene adapted consortium accelerated dechlorination which occurred at a calculated rate 

of 0.29 mg/L/day. At hexachlorobenzene concentrations of 2, 5, and 10 mg/L, complete dechlorination 

occurred within 6 days and at the 50 mg/L concentration, dechlorination occurred in 8 days. Optimal 

dechlorination occurred at a rate of 0.29 mg/L/day, pH of 7.0, and 30 EC. According to Yuan et al. 

(1999), dechlorination occurred via the following path: hexachlorobenzene 6 pentachloro-benzene 6

1,2,3,4-tetrachlorobenzene + 1,2,3,5-tetrachlorobenzene 6 1,2,4-trichlorobenzene + 1,2,3-trichlorobenzene + 1,3,5-trichlorobenzene 6 1,2-dichlorobenzene + 1,4-dichlorobenzene. From this and other 
HEXACHLOROBENZENE 216 

6. POTENTIAL FOR HUMAN EXPOSURE 

studies, it is clear that in a time frame of days to years, anaerobic biodegradation may remove hexachlorobenzene from soils. 

The Chemicals Inspection and Testing Institute (1992) of Japan conducted an aerobic sludge study to test 

hexachlorobenzene biodegradation. Hexachlorobenzene, present at 100 mg/L, reached 0% of its 

theoretical biological oxygen demand (BOD) in 2 weeks using an activated sludge inoculum at 30 mg/L 

and the Japanese Ministry of International Trade and Industry (MITI) test; thus, aerobic degradation is not 

an important fate process. 

Likewise, in areas of the Great Lakes region with a long history of hexachlorobenzene contaminated 

waste water discharges affecting aquatic sediments, the concentrations of hexachlorobenzene in the 

sediments can be significant (see Section 6.4.2). Susarla et al. (1997) examined the transformation of 

hexachlorobenzene in fresh water lake (Lake Kasumigaura, Japan) sediments under anaerobic conditions. 

Dechlorination occurred after a 4-day lag phase and was complete in 32 days. The calculated rate of 

dechlorination was 0.110/day. Hexachlorobenzene transformation pathway under sulfidogenic conditions 

resulted in hexachlorobenzene 6 pentachlorobenzene 6 1,2,3,5-tetrachlorobenzene 6 1,3,5-trichlorobenzene 6 1,3-dichlorobenzene. Under methanogenic conditions the pathway was as follows: 

hexachlorobenzene 6 pentachlorobenzene 6 1,2,3,4-tetrachlorobenzene 6 1,2,4-trichlorobenzene 6

1,4-dichlorobenzene. After almost a year, 98% of the hexachlorobenzene was dechlorinated to 1,3- and 

1,4-dichlorobenzene (Susarla et al. 1997). In another experiment, dechlorination of hexachlorobenzene in 

an estuary sediment collected from the mouth of Tsurumi river occurred at a rate of 0.0256/day with a 

half-life of 27.1 days (Masunaga et al. 1996). Thus, aquatic sediment degradation of hexachlorobenzene 

occurs in a month to a year. 

6.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT 

Reliable evaluation of the potential for human exposure to hexachlorobenzene depends in part on the 

reliability of supporting analytical data from environmental samples and biological specimens. In 

reviewing data on hexachlorobenzene levels monitored in the environment, it should also be noted that 

the amount of chemical identified analytically is not necessarily equivalent to the amount that is 

bioavailable. 
HEXACHLOROBENZENE 217 

6. POTENTIAL FOR HUMAN EXPOSURE 

6.4.1 Air 

Hexachlorobenzene is moderately volatile, but is usually not detected in ambient air samples except at 

very low concentrations. Ambient air sampling was conducted by the EPA at selected locations in the 

United States from 1976 to 1979. In 1976, 49% of the 43 composite samples collected in four locations 

contained detectable concentrations of hexachlorobenzene, with a mean value of 0.1 ng/m3

. In 1977, 12% 

of the 34 samples collected in three locations contained hexachlorobenzene, with a mean concentration of 

1.5 ng/m3

. In 1978, the compound was not detected in any of the 33 samples collected at three locations 

and in 1979, hexachlorobenzene was detected in 31% of the 89 samples collected in eight locations, with 

a mean of 0.5 ng/m3

 (detection limit: 0.1 ng/m3) (EPA 1985g). Eisenreich et al. (1981) reported 

atmospheric concentrations of hexachlorobenzene in the Great Lakes region ranging from 0.09 to 

0.28 ng/m3

. Results of airborne samples collected between 1990 and 1993 from the Great Lakes region 

by the Integrated Atmospheric Deposition Network are provided by Hoff et al. (1996). The annual mean 

gas-phase and particulate-phase concentrations of hexachlorobenzene were 98 and 0.2 pg/m3

, 

respectively, in samples from Lake Superior near Eagle Harbor, Michigan, 120 and 0.1 pg/m3

, 

respectively, in samples from Lake Michigan near Sleeping Bear Dunes, Michigan, 80 and 0.2 pg/m3

, 

respectively, in samples from Lake Erie near Sturgeon Point, New York, and 130 and <0.1 pg/m3

, 

respectively, in samples from Lake Ontario near Point Petre, Ontario. From July 1988 to September 

1989, 143 air samples were collected at Egbert, Ontario, Canada and were analyzed for PCB and 

organochlorine concentrations. Hexachlorobenzene was detected at concentrations ranging from a 

minimum of 0.04 pg/m3

 (0.00004 ng/m3) to a maximum of 640 pg/m3 (0.64 ng/m3) (annual mean 

>54 pg/m3

 (0.054 ng/m3) (Hoff et al. 1992). Inhalation exposure was estimated to be 0.3 ng/m3 in urban 

air (Burton and Bennett 1987). Hexachlorobenzene measured in air in Villeroy, Quebec in 1992 found 

mean and median concentrations of 36.68 (0.03668) and 30.94 pg/m3

 (0.03094 ng/m3), respectively, from 

56 air samples (Poissant et al. 1997). A meteorological station located in a semirural area outside 

Lancaster, England was the site of air samples. Four air samples per day (taken at 6-hour intervals) were 

taken for 7 days. The minimum, maximum, and mean concentrations of hexachlorobenzene in these 

samples were <28.8, 76.1, and 39.3 pg/m3

, respectively. The authors found an absence of a cycle in the 

concentrations of hexachlorobenzene and concluded that the compound was breaking through the 

polyurethane foam plugs due to its relatively high vapor pressure (Lee et al. 2000). 

Hexachlorobenzene air concentrations have also been measured in urban and rural areas in France. 

Atmospheric fallout from the urban area, Paris, and the rural area, La Ferté-sous-Jouarre, was collected in 

raw form as bulk precipitation. Hexachlorobenzene concentration in rural fallout measured in 
HEXACHLOROBENZENE 218 

6. POTENTIAL FOR HUMAN EXPOSURE 

February–July 1992 and January –September 1993 ranged from 2.5 to 4.5 ng/L and from 0.3 to 4 ng/L, 

respectively. For the same time periods, urban fallout measured 1.8–17 and 0.3–5.6 ng/L, respectively 

(Chevreuil et al. 1996). The mean concentrations of hexachlorobenzene in precipitation samples 

collected in the Great Lakes region from 1986 to 1991 ranged from 0.145 ng/L (ppt) at Sibley Park on 

Lake Superior, to 0.108 ng/L (ppt) at Pelee Island in Lake Erie, to 0.174 ng/L (ppt) at Wolfe Island in 

Lake Ontario (Chan et al. 1994). The mean and median concentrations of hexachlorobenzene from eight 

precipitation samples collected from Villeroy, Quebec in 1992 were 0.04 and 0.05 ng/L, respectively 

(Poissant et al. 1997). Precipitation samples collected between 1990 and 1993 from the Great Lakes 

region by the Integrated Atmospheric Deposition Network were analyzed. The annual mean 

concentration of hexachlorobenzene in these precipitation samples were 0.1 ng/L in samples from Lake 

Superior near Eagle Harbor, Michigan, 0.06 ng/L in samples from Lake Michigan near Sleeping Bear 

Dunes, Michigan, 0.04 ng/L in samples from Lake Erie near Sturgeon Point, New York, and 0.3 ng/L in 

samples from Lake Ontario near Point Petre, Ontario (Hoff et al. 1996). The concentrations of hexachlorobenzene in air, particulate matter, and rain from Galveston Bay, Texas were 87.3±103.3, 0.4±0.5, 

and 42.5 pg/m3

 (not detected–48.1 pg/m3), respectively, between 1995 and 1996 (Park et al. 2001). The 

median levels of hexachlorobenzene in ambient air samples collected in Zagreb, Croatia in 1997 were 

29 pg/m3

 (range, 0.5–49 pg/m3) in the northern residential region of Ksaverska and 31 pg/m3 (range, 

15–61 pg/m3

) in the southern region near a landfill (Romaníc and Krauthacker 2000). The average 

concentration of hexachlorobenzene in air at Lake Malawi, in southeast Africa, from February 1997 to 

May 1998, was 11±7.5 pg/m3

 (Karlsson et al. 2000). 

Nonoccupational exposure to hexachlorobenzene for residents of two U.S. cities (Jacksonville, Florida 

and Springfield, Massachusetts) was studied over three seasons: summer 1986, spring 1987, and winter 

1988 (Whitmore et al. 1994). The study focused primarily on inhalation exposures with primary 

environmental monitoring consisting of 24-hour indoor and outdoor air. For the Jacksonville, Florida, 

population, the estimated mean hexachlorobenzene concentration ranged from 0.3 to 1.3 ng/m3

 for indoor 

air and from not detected to 0.2 ng/m3

 for outdoor air. For the Springfield, Massachusetts population, 

mean exposures were much less. The estimated hexachlorobenzene concentrations ranged from not 

detected to 0.1 ng/m3

 for indoor air and no detectable concentrations of hexachlorobenzene for outdoor 

air. 

Extremely high concentrations of hexachlorobenzene in air have been detected in areas close to 

production and disposal sites in both outdoor and indoor air. Mann et al. (1974) measured hexachlorobenzene concentrations ranging from 70 to 23,296 ng/m3

 near chlorinated solvent and pesticide 
HEXACHLOROBENZENE 219 

6. POTENTIAL FOR HUMAN EXPOSURE 

manufacturing facilities; air levels near a chemical waste landfill were as high as 16,000 ng/m3 (EPA 

1975b). Hexachlorobenzene has been detected at 11,000 ng/m3

 in flue gas effluents from a municipal 

refuse-fired steam boiler in Virginia, and at 9.5 ng/m3

 in flue gas effluents from a refuse-derived fuel fired 

power plant in Ohio (Tiernan et al. 1985). Air concentrations of hexachlorobenzene inside industrial 

plants can be as high as 150,000 ng/m3

 (Currier et al. 1980); air concentrations inside a pesticide 

production facility were measured at 22,000 ng/m3

 (Davis and Morgan 1986). 

6.4.2 Water 

Drinking water in three cities in the Lake Ontario vicinity contained hexachlorobenzene ranging from 

0.06 to 0.2 ppt (mean of 0.1 ppb), about the same as water from the lake (Oliver and Nichol 1982a). 

Hexachlorobenzene was detected in ambient water samples from two of the Great Lakes and their 

tributary rivers. Mean concentrations of hexachlorobenzene in Lake Ontario, Lake Huron, and the Grand 

River were 0.06 ppt (range, 0.02–0.1 ppt), 0.04 ppt (range, 0.02–0.1 ppt), and 0.06 ppt (range, 

0.02–0.1 ppt), respectively. In the Niagara River, concentrations of 0.02–17 ppt were detected with the 

highest value measured downstream of a waste disposal site (Oliver and Nichol 1982a). Widely varying 

measurements in this river may be attributed to the fact that measurements were near the limit of the 

analytical detection limits. More recently, hexachlorobenzene was detected in 42% of whole water 

samples (dissolved plus particulate phases) collected during 1991 in the St. Lawrence River and several of 

its tributaries. Hexachlorobenzene concentrations detected in the St. Lawrence River ranged from not 

detected to 0.09 ng/L (mean 0.01 ng/L [ppt]) (Quemerais et al. 1994). Hoff et al. (1996) obtained and 

presented water concentration data from a 1992 sampling study. Hexachlorobenzene concentrations in 

Lakes Superior, Michigan, Huron, Erie, and Ontario were 0.01, 0.014, 0.007, 0.014, and 0.045 ng/L, 

respectively. In remote European mountain lake waters, the concentrations of hexachlorobenzene were 

8.4±11 pg/L at Redó, Spain, 4.0±1.8 pg/L at Gossenkölle, Austria, and 6.2±1.0 pg/L at Øvre Neådalsvatn, 

Norway (Vilanova et al. 2001). The concentrations of hexachlorobenzene in microlayer and subsurface 

Mediterranean seawater off the coast of Alexandria, Egypt were on average, 27.3±17 and 12±6.9 ng/L, 

respectively (Abd-Allah 1999). 

A study was conducted from 1974 to 1975 to collect and analyze surface water samples from sites of 

known hexachlorobenzene contamination along the Mississippi River near an industrial area in Geismar, 

Louisiana (EPA 1976a). The maximum hexachlorobenzene concentration detected in water was 
HEXACHLOROBENZENE 220 

6. POTENTIAL FOR HUMAN EXPOSURE 

90.3 ppb. A concentration of 2 ppb has been measured in the Mississippi River near Baton Rouge, 

Louisiana (Laska et al. 1976). 

Industrial waste water samples contained hexachlorobenzene levels as high as 300 ppb (EPA 1976b; 

Schmitt et al. 1990). Effluent concentrations of hexachlorobenzene from four Canadian plants in the 

Great Lakes region ranged from 0.001 to 0.002 ppb (0.0015 ppb mean) (Oliver and Nichol 1982a). 

Hexachlorobenzene concentrations were measured in water at an uncontrolled hazardous waste site near 

Bayou Baton Rouge, Louisiana (Davis and Morgan 1986). Surface water samples collected from a 

containment pond used for disposal of wastes from both rubber production and manufacture of 

chlorinated organics at the site contained up to 8,100,000 ppb hexachlorobenzene. 

6.4.3 Sediment and Soil 

Mean concentrations of hexachlorobenzene in lake sediments in the Great Lakes ranged from 0.2 to 

97 ppb with the highest values measured in Lake Ontario. Deeper sediment layers (1–2 cm) had even 

higher concentrations of hexachlorobenzene (460 ppb) than surface (0–1 cm) samples (270 ppb), with the 

peak value corresponding to deposition in the years 1971–1976 declining to 270 ppb in 1976–1980 

(Oliver and Nichol 1982a). In 1992, 2 cm deep bed sediment samples were collected from the South 

Platte River at Henderson, Colorado and Cache La Poudre River near Greeley, Colorado. The sediment 

contained 1.5 and <1 μg/kg, dry weight, hexachlorobenzene, respectively. The authors concluded that 

this concentration was correlated to the hexachlorobenzene concentrations found in urban and agricultural 

lands in the South Platte River Basin (Tate and Heiny 1996). Outside of the United States, 12 sediment 

samples were collected in June 1993 near known discharges from municipalities and industries from Lake 

Ladoga, Russia. Hexachlorobenzene concentrations were 3.58 and 14.6 ng/g in 2 out of the 12 samples, 

and was not detected in the remaining 10 samples (Ristola et al. 1996). 

In 1972, hexachlorobenzene levels in soil were detected in 11 of 1,485 agricultural sites (0.7%) in 

37 states, ranging from 0.01 to 0.44 mg/kg (10–440 ppb). In 1973, only 1 of 1,470 sites (0.1%) contained 

hexachlorobenzene at 10 ppb, while in a 1976 study of 11 states, only 2 of 391 sites (0.5%) contained 

hexachlorobenzene at concentrations ranging from 10 to 20 ppb (EPA 1985g). The majority of these 

agricultural sites with hexachlorobenzene detections were under cultivation with wheat or other seed 

grains (e.g., barley, oats, rye), which is consistent with the registered agricultural uses of hexachlorobenzene at that time. The authors concluded that the occurrence of hexachlorobenzene residues in 
HEXACHLOROBENZENE 221 

6. POTENTIAL FOR HUMAN EXPOSURE 

agricultural soils was associated with hexachlorobenzene's registered pesticide uses rather than general 

environmental or industrial contamination. 

In a study of 40 urban areas sampled during the 1970s, 7 sites (17.5%) contained detectable hexachlorobenzene concentration in soil ranging from 0.01 to 0.59 mg/kg (10–590 ppb) (EPA 1985g). In the levees 

of the Mississippi River near Baton Rouge, where the river water contained 2 ppb hexachlorobenzene, the 

soil contained 167 ppb (Laska et al. 1976). In contrast to agricultural soils, hexachlorobenzene residues 

in urban soils resulted from releases during the manufacture, use, or disposal of hexachlorobenzene or 

hexachlorobenzene-containing wastes rather than the use of hexachlorobenzene as a pesticide (EPA 

1985g). 

Hexachlorobenzene concentrations were measured in soil and sediment at several uncontrolled hazardous 

waste sites in several states (Davis and Morgan 1986). Hexachlorobenzene concentrations of up to 

20,000 ppb were measured in soil at a scenic highway site near Bayou Baton Rouge, Louisiana, while 

concentrations in sediment of 39,500 ppb were measured from a bayou bank downstream of the site. Soil 

cores from a monitoring well (25–27 feet deep) were as high as 400,000 ppb, and as high as 90,000 ppb 

in soils collected at 40–41 feet deep (Davis and Morgan 1986). Soil and sediment collected from a 

disposal site near Sorrento, Louisiana contained 62,000 and 130,000 ppb hexachlorobenzene, 

respectively. Soil collected at a Crystal City, Texas pesticide disposal site was found to contain 

20,000 ppb hexachlorobenzene. A maximum hexachlorobenzene concentration detected in soil at an 

industrial site of known contamination near Geismar, Louisiana was 53,130 ppb (Laseter et al. 1976). 

Sediment concentrations of hexachlorobenzene vary widely from relatively unpolluted areas to those 

areas used extensively for disposal of hexachlorobenzene-containing wastes. Sediment hexachlorobenzene concentrations from San Luis Pass, located near industrial areas of West Galveston Bay, Texas 

ranged from 0.05 to 1.5 ppb (dry weight) with a mean of 0.49 ppb (Murray et al. 1981). 

Hexachlorobenzene concentrations in marine sediment samples collected from an industrialized area of 

the harbor in Portland, Maine ranged from <0.03 to 0.37 ppb (mean 0.14 ppb) (Ray et al. 1983). 

Concentrations of hexachlorobenzene in sediment from the Niagara River watershed in the vicinity of 

several hazardous waste disposal areas ranged from 8,000 to 30,000 ppb (Elder et al. 1981). The average 

concentration of hexachlorobenzene in surficial sediments of the Kaohsiung coast (southwestern Taiwan), 

which receives wastewater from the largest industrial city in Taiwan (Kaohsiung City), ranged from 1.7 to 

24.7 ng/g (Lee et al. 2000a). 
HEXACHLOROBENZENE 222 

6. POTENTIAL FOR HUMAN EXPOSURE 

6.4.4 Other Environmental Media 

Concentrations of hexachlorobenzene have been detected in several species of fish and shellfish. Hexachlorobenzene concentrations were determined for several species of marine organisms collected from 

San Luis Pass near Galveston Bay, Texas (Murray et al. 1981). Mean concentrations of hexachlorobenzene at were 0.49 ppb wet weight for flounder (species unspecified), 0.65 ppb for longnose killifish 

(Fundulus similis), 0.88 ppb for brown shrimp (Penaeus aztecus), 9.6 ppb for blue crab (Callinectes 

sapidus), and 0.71 ppb for the dwarf squid (Lollingnucula brevis). Oysters (Crassotrea virginica) 

collected at the lower end of the Houston Ship Channel were found to contain hexachlorobenzene 

concentrations ranging from 0.31 to 1.41 ppb with a mean of 0.63±0.39 ppb (Murray et al. 1980). The 

Sheboygan River in Wisconsin is another area of concern for contamination of organochlorine pesticides 

in fish due to the existence of wetlands, urban and developed land, woodland, and agricultural land 

surrounding this river. In addition, the lower segment of the river has a history of shipping, industrial and 

municipal activities, and dredging, including the existence of a landfill designated as a federal superfund 

site. Schrank et al. (1997) collected white suckers (Catostomus commersoni) from two sites of the 

Sheboygan River; one site was 1 to 2 km from the mouth of the river and the other was 50 km from the 

mouth of the river, which served as the reference site. The fish collected from both sites contained less 

than detectable residues of hexachlorobenzene along with other organochlorine compounds, thus 

minimizing the risk of exposure to hexachlorobenzene from this river. 

DeVault (1985) reported concentrations of hexachlorobenzene in whole fish composites collected from 

the Great Lakes during 1980 and 1981 ranged from <0.002 to 3.47 mg/kg (<2–347 ppb). Hexachlorobenzene concentrations in fresh water trout (4–6+ years old) from the Great Lakes region ranged from 

8 to 127 ppb with the highest concentration found in a fish collected near the discharge of the Niagara 

river into Lake Ontario (Oliver and Nichol 1982a). In another study of Great Lakes fish species, 

Newsome and Andrews (1993) reported hexachlorobenzene concentrations in fish fillet composites 

ranged from 0.22 ng/g (ppb) in bullhead to 9.05 ng/g (ppb) wet weight in trout in lake areas with open 

fisheries. Zabik et al. (1995) reported that skin-off processing and selected cooking methods reduced 

hexachlorobenzene residues in chinook salmon and carp harvested from the Great Lakes. Concentrations 

of hexachlorobenzene averaged 0.017 ppm (17 ppb) and 0.011 ppm (11 ppb) (wet weight) in raw and 

cooked salmon fillets, respectively, and averaged 0.005 ppm (5 ppb) and 0.003 ppm (3 ppb) in skin-on 

and skin-off fillets, respectively. The average percentage loss of hexachlorobenzene from chinook 

salmon fillets by baking, charbroiling, and canning was 40%. Losses of hexachlorobenzene residues from 

carp fillets were slightly greater than 40% (Zabik et al. 1995). Walleye, siscowet, carp, and whitefish 
HEXACHLOROBENZENE 223 

6. POTENTIAL FOR HUMAN EXPOSURE 

were collected for organochlorine pesticides analysis from Lake Superior along the Apostle Islands region 

during 1991 and 1992. Walleye and carp had hexachlorobenzene concentrations below the limit of 

quantification, while siscowet and whitefish measured concentrations of 3.2 and 2.8 ng/g wet weight of 

tissue, respectively (Gerstenberger et al. 1997). Organisms sampled during the summer of 1994, from the 

Keweenaw Peninsula of Lake Superior, contained measured hexachlorobenzene concentrations ranging 

from 0.8 to 1.8 ng/g wet weight in smelts, 3.0–4.3 ng/g wet weight in herrings, 4.7–8.4 ng/g wet weight in 

bloaters, 1.1–4.1 ng/g wet weight sculpins, <0.1–0.2 ng/g wet weight in mysis, 0.8–1.4 ng/g wet weight in 

limnocalanus, 0.8 ng/g wet weight in amphipods, and 1.7–3.1 ng/g wet weight in lake trout (Kuchklick 

and Baker 1998). Grayling and lake trout were collected from four lakes (Feniak, Desperation, Schrader, 

and Elusive) from the foothills of the Brooks Range, Alaska. All lakes were oligotrophic, and the pH was 

neutral to slightly alkaline. Fifty-six grayling liver samples and 39 grayling muscle samples were 

analyzed for hexachlorobenzene concentration, and mean and median values were derived. In the 

56 grayling liver samples, the mean and median concentrations were 0.65 and 0.48 ng/g dry weight, 

respectively. The 39 grayling muscle samples had mean and median concentrations of 0.33 and 0.22 ng/g 

dry weight, respectively. In lake trout, the mean and median concentrations in 33 liver samples were 

1.15 and 0.87 ng/g dry weight, respectively, and in 34 muscle samples were 0.46 and 0.26 ng/g dry 

weight, respectively (Allen-Gil et al. 1997a). Hexachlorobenzene concentrations in sea organisms from 

the Barents Sea were as follows (units=ng/g lipid weight): copepods (13.5), euphausids (16.5), amphipods 

(19.5), polar cod (39±1.7), and cod (65±7.7) (Borgå et al. 2001). In 1991, the concentrations of hexachlorobenzene in amphipods, isopods, and sculpins from the Bothnian Bay (Baltic Sea) were 340 (n=3), 

370 (n=5), and 37 (n=3) ng/g dry weight, respectively (Strandberg et al. 2000). 

The bioaccumulative tendencies of hexachlorobenzene have made it a candidate for monitoring in the 

U.S. Fish and Wildlife Service National Pesticide Monitoring Program (Schmitt et al. 1990) and the 

National Study of Chemical Residues in Fish which was started in 1986 (NSCRF) (EPA 1992b). 

Maximum hexachlorobenzene tissue concentrations (wet weight) detected in whole fish were 700, 130, 

120, and 410 ppb in the 1976–1977, 1978–1979, 1980–1981, and 1984 sampling years, respectively. The 

geometric mean tissue concentration was 10 ppb for 1976–1977 and <10 ppb for all other sampling years 

(Schmitt et al. 1990). The highest hexachlorobenzene concentrations in the 1984 sampling period 

(410 ppb) were detected in whole fish from the Tombigbee River, Alabama in the vicinity of a pesticide 

production facility where concentrations during all sampling years had been the highest nationally. The 

most recent national results from the ongoing study conducted by NSCRF show that hexachlorobenzene 

was detected at 46% of the 362 sites surveyed for fish tissue analysis. The mean hexachlorobenzene 

concentration for fish tissue samples analyzed in this program was 5.8 ppb. The five sites with the 
HEXACHLOROBENZENE 224 

6. POTENTIAL FOR HUMAN EXPOSURE 

highest concentrations are listed in Table 6-3 (EPA 1992b). The Freeport, Texas site is near a pesticide 

plant and the other four sites are close to a variety of chemical manufacturing plants. The Calcasieu 

River, Louisiana site is close to a Superfund site involving a variety of organic solvents (EPA 1992b). 

Hazardous waste dumping during the early 1940s and 1950s contaminated the Devil’s Swamp, Louisiana 

with chlorinated hydrocarbons, which has greatly affected fish species. As fish is an important food 
HEXACHLOROBENZENE 225 

6. POTENTIAL FOR HUMAN EXPOSURE 

Table 6-3. Sites with the Five Highest Concentrations of 

Hexachlorobenzene in Fisha
 

Whole body 

hexachlorobenzene Type 

concentration of 

(ppb; wet weight) sample (fish) Location 

913 Sea catfish Brazos River, Freeport, Texas 

202 Catfish Bayou D'Inde, Sulfur, Louisiana 

93.7	 Carp Mississippi River, St. Francisville,
 

Louisiana
 

85.5	 White sucker Quinipiac River, North Haven,
 

Connecticut
 

75 Sea catfish	 Calcasieu River, Moss Lake, Louisiana 

a

From the EPA 1992b National Study of Chemical Residues in Fish 
HEXACHLOROBENZENE 226 

6. POTENTIAL FOR HUMAN EXPOSURE 

source for the community surrounding this area, concentrations of contaminants are of great concern. 

Levels of hexachlorobenzene in various fish species collected from Devil’s Swamp, Louisiana, was 

compared to those from a control site, Tunica’s Swamp, Louisiana. Mean hexachlorobenzene 

concentrations calculated from concentrations of 13 different edible fish species tissues were 23.65 ng/g 

compared to 2.0 ng/g calculated from 10 different edible fish species from Tunica’s Swamp (Tchounwou 

et al. 1998). 

Hexachlorobenzene has been detected in tissues of various wildlife species throughout North America, 

but especially in wildlife indigenous to the Great Lakes region. Swift et al. (1993) reported mean 

concentrations of hexachlorobenzene of 0.02 ppm (20 ppb) wet weight (0.04 ppm [40 ppb] lipid weight 

basis) and 0 ppm wet weight (0.07 ppm [70 ppb] lipid weight basis) in mesenteric and subcutaneous fat 

and breast tissue, respectively, of common goldeneye waterfowl wintering in New York state. The 

detection limit in this study was 0.002 ppm. Gebauer and Weseloh (1993) reported that the geometric 

mean hexachlorobenzene concentrations of 0.4 and 0.9 μg/kg (ppb) in muscle tissue in mallard ducks 

using a contaminated sediment site, and sewage lagoon site, respectively, as a resting and feeding area 

were significantly greater than levels found in ducks using a natural marsh area. Foley (1992) reported 

hexachlorobenzene residues in muscle tissues of several species of ducks and geese collected in New 

York State in 1983–1984. Statewide residues were 64 ppb (wet weight) for buffleheads, 49 ppb for 

scaups, 26 ppb for mallards, 20 ppb for black ducks, 6 ppb for wood ducks, and 11 ppb for Canada geese. 

Adult sea otters that had died along the coast of California were collected by the U.S. Fish and Wildlife 

Service and the California Department of Fish and Game. Hexachlorobenzene concentrations in liver, 

kidney, and brain tissues were 0.74–8, 0.28–2.6, and 0.28–0.74 ng/g wet weight, respectively (Nakata et 

al. 1998). The mean hexachlorobenzene concentration for 207 wild mink liver tissue samples, collected 

from 1991 to 1995 in the Northwest Territories, Canada from seven mink communities, ranged from 

0.21 to 0.67 ng/g wet weight (Poole et al. 1998). Snail composite samples, without shells, were collected 

from two lakes (Feniak and Elusive) from the foothills of the Brooks Range, Alaska. All lakes were 

oligotrophic, and the pH was neutral to slightly alkaline. The mean concentration among these six 

samples was 0.15 ng/g dry weight with a median of 0.10 ng/g dry weight (Allen-Gil et al. 1997a). 

Hexachlorobenzene has also been detected in the eggs of various wildlife species in the Great Lakes 

region and Canada. Yamashita et al. (1992) reported hexachlorobenzene residues ranges of 8–36 ng/g 

(ppb) and 18–26 ng/g (ppb) on a wet weight basis in the eggs of the double-crested cormorant and the 

Caspian tern, respectively, collected during 1988 from the Great Lakes region. Somers et al. (1993) 

reported geometric mean concentrations of 0.013 μg/g (13 ppb) (wet weight) of hexachlorobenzene in 
HEXACHLOROBENZENE 227 

6. POTENTIAL FOR HUMAN EXPOSURE 

double-crested cormorant eggs collected in southern Alberta, Canada. Elliott and Martin (1994) reported 

mean hexachlorobenzene concentrations in sharp-shinned hawk eggs in south central Ontario ranging 

from 0.010 to 0.051 mg/kg (10–51 ppb) from 1986 to 1989. Hexachlorobenzene concentrations in 

Cooper's hawk eggs ranged from 0.005 to 0.012 mg/kg (5–12 ppb) during the same period. Cobb et al. 

(1994) reported mean residues of 18.0 ng/g (18 ppb) in the chorio-allantoic membranes removed from 

great blue heron eggs collected from Puget Sound, Washington. Jarman et al. (1996) conducted an 

experiment with prairie falcon eggs that were collected from eyries in northern and central California 

between 1989 and 1991. Addled and unhatched eggs were frozen until chemical analysis. The following 

are the geometric mean concentrations (in mg/kg wet weight) of hexachlorobenzene at their respective 

sampling sites: Frog/Hand Nest 800; Pig Cyn 17; Crowley Tower 11; Willow Sp. 8.0; Goat Rock 10; Mt. 

Dome 20; and Mt. Diablo 81. Mean hexachlorobenzene residues in peregrine falcon eggs from Rankin 

Inlet (Hudson Bay, Canada) were 0.03 μg/g wet weight (n=2; range 0–0.15 μg/g wet weight) between 

1982 and 1986, and 0.030 μg/g wet weight (n=20; range 0–0.165 μg/g wet weight) between 1991 and 

1994 (Braune et al. 1999). 

Hexachlorobenzene residues were also detected in snapping turtle eggs collected from a wetland area on 

Lake Ontario. Residues ranged from 43.9, 16.6, and 20.9 ng/g (ppb) (wet weight) to 494.7, 282.1, and 

262.2 ng/g (ppb) (lipid weight) for the first five eggs, a composite of five eggs, and the last five eggs, 

respectively (Bishop et al. 1995). Bishop et al. (1996) conducted another study with snapping turtle eggs 

that were collected from nests at five locations from the Great Lakes Basin in 1990–1991. Thee eggs 

were analyzed for hexachlorobenzene and the results were compared to data collected from the same sites 

in the years 1981, 1984, 1988, and 1989. Residue concentrations for all locations and their respective 

years are listed in Table 6-4. Based on the results, the hexachlorobenzene concentration was the highest 

in eggs from Cootes Paradise and lowest in eggs from Algonquin Park. Although the concentration of 

hexachlorobenzene at Cootes Paradise has been continuously declining since 1984, residues in eggs from 

this site are consistently higher than those from the other sites, and are at least 3 times higher than eggs 

from the next most contaminated site, Lynde Creek. On the whole, the hexachlorobenzene mean 

concentration from the five sites in the Great Lakes Basin showed a decrease from the year 1984 to 1990 

(Bishop et al. 1996). 

Langlois and Langis (1995) reported that the concentration of hexachlorobenzene in the blubber of beluga 

whales from the St. Lawrence Estuary to Northern Quebec Province ranged from 0.22 to 
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6. POTENTIAL FOR HUMAN EXPOSURE 

Table 6-4. Residue Concentrations (ng/g) in Snapping Turtle Eggs 

from Five Locations Collected in 1981–1991 

Year 

Sites 1981 1984 1988 1989 1990/1991 

Algonquin 20 NA 10 3 NA 

Park 

Cranberry NA NA 40 30 NA 

Park 

Lynde Creek NA 70 80 60 80 

Cootes NA 350 300 170 170 

Paradise 

Big Creek 50 NA NA 40 NA 

Marsh 

Rondeau Park NA 60 NA 20 NA 

NA = Snapping turtle eggs were not available or were not collected during those years. 

Source: Bishop et al. 1996 
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6. POTENTIAL FOR HUMAN EXPOSURE 

0.93 mg/kg (220–930 ppb) wet weight. Consumption of blubber and organ meats from these whales by 

native peoples may constitute a potential health risk if these tissues are a significant part of the diet 

(Langlois and Langis 1995). In 1997, Gauthier et al. (1997) analyzed blubber biopsies from 

Northwestern Atlantic Balaenopterid whales summering in the Gulf of St. Lawrence. Samples were 

collected in the summer and fall of 1991 and 1992 from 21 minke, 15 fin, 6 blue, and 8 humpback 

whales. Hexachlorobenzene concentrations in the blubber of these whales were 101, 96, 110, and 

177 ng/g lipid, respectively, and in all species, the concentrations were higher in males (140 ng/g lipid) 

than in females (103 ng/g lipid). 

Becker et al. (1997) analyzed the blubber of 7 pilot whales, 5 harbor porpoises, 12 beluga whales, 

2 northern fur seals, and 2 ringed seals that were obtained from the U.S. National Biomonitoring 

Specimen Bank. Hexachlorobenzene concentration ranges were 43–465 ng/g wet weight for pilot whales, 

223–1,070 ng/g wet weight for harbor porpoises, 81.9–952 ng/g wet weight for beluga whales, 

138–741 ng/g wet weight for northern fur seals, and 125–156 ng/g wet weight for ringed seals (Becker et 

al. 1997). Elkin and Bethke (1995) reported that hexachlorobenzene was the most predominant organochlorine residue present in tissues of the caribou collected in the Northwest Territory of Canada. 

Residues ranged from a lipid corrected mean of 32.83 ng/g (ppb) in fat from Bathurst caribou to 

129.41 ng/g (ppb) in Lake Harbor animals (Baffin Island). Consumption of meat and organs from these 

range animals by native peoples, including native American populations of Inuits in Alaska, may 

constitute a potential human health risk if these tissues are a significant part of the diet. 

The Florida Department of Environmental Protection and the Florida Marine Research Facility in St. 

Petersburg, Florida maintain archives of tissues obtained from dead Florida Manatees (Trichechus 

manatus latirostris). In 1996, Ames and Van Vleet (1996) obtained and analyzed 19 manatee samples 

from the Florida EPA. Of these samples, hexachlorobenzene was found at concentrations of 0.038 and 

0.085 μg/g in one blubber and one liver sample, respectively. The authors did not find a correlation 

between the contamination in manatees by hexachlorobenzene and other pesticides and the location in 

which the manatees were found dead; thus, the authors concluded that these manatees must have been 

contaminated elsewhere, especially since manatees are known to travel long distances. 

Corsolini et al. (1999) analyzed chlorinated hydrocarbon concentrations in muscle and fat samples of the 

red fox collected from Sienna, Italy in 1992. Hexachlorobenzene mean concentrations in muscle and fat 

were 0.47 and 0.23 μg/g lipid basis, respectively, and were the lowest of all of the chlorinated 

hydrocarbons that were tested (Corsolini et al. 1999). 
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6. POTENTIAL FOR HUMAN EXPOSURE 

Sitarska et al. (1995) collected tissue samples from 18 cows just after slaughtering, always from the same 

parts of the studied organs. Hexachlorobenzene mean concentrations were 1.394 μg/kg wet mass in the 

ovaries, 1.061 μg/kg wet mass in the mammary glands, and 0.550 μg/kg wet mass in the liver. 

Beyer (1996) subjected earthworms (Lubricous terrestris) to hexachlorobenzene in artificial soil. Over 

the course of 28 weeks, hexachlorobenzene concentrations in these earthworms ranged from 1.8 to 3 ppm. 

Beyer (1996) also conducted three 8-week experiments in which earthworm survival rates were 97, 99, 

and 100%, respectively. 

Hexachlorobenzene was detected in composited milk samples collected through the U.S. Pasteurized Milk 

Network during 1990–1991 (Trotter and Dickerson 1993). The milk samples were collected at 

approximately 63 sites located in the United States, Puerto Rico, and the Panama Canal Zone. Hexachlorobenzene was detected in trace amounts in one sample collected in each of seven metropolitan areas 

(Cristobal, Panama Canal Zone; Kansas City, Missouri; Los Angeles, California; Memphis, Tennessee; 

Portland, Oregon; Spokane, Washington; Wichita, Kansas) and was detected at 0.001 ppm (1 ppb) in one 

sample from Rapid City, South Dakota. 

Pesticide residue data from the FDA Adult Total Diet Study conducted from 1980 to 1982 were evaluated 

by Gartrell et al. (1986). These authors reported that hexachlorobenzene was detected in a wide variety of 

domestic foods: dairy products, meat, fish and poultry, oils and fats, and sugar and adjuncts. The highest 

mean concentrations of hexachlorobenzene were detected in oils and fats (0.9 ppb) and in meat, poultry, 

and fish (0.2 ppb). Concentrations of hexachlorobenzene in ready-to-eat foods were monitored for 

10 years from 1982–1991 through the FDA's Revised Market Basket Survey. Hexachlorobenzene was 

detected in 618 samples of 81 different foods at a mean concentration of 0.0006 μg/g (0.6 ppb) (KAN-DO 

Office and Pesticide Teams 1995). In food composites from six Canadian cities, the mean concentration 

of hexachlorobenzene in positive samples (4.8% of 913 total analysis) was 0.5 ng/g (Newsome et al. 

2000). The U.S. Food and Drug Administration monitored domestic and imported apples and rice by 

collecting random samples for a period of 12 months beginning in October, 1993. Hexachlorobenzene 

was not determined to be in violation according to the concentration limits set for this compound in any 

of the domestic and imported apple and rice samples; however, it was found in 1 out of 612 imported rice 

samples (0.02 ppm), but this concentration does not violate any limit set by the EPA (Roy et al. 1997). 

The concentrations and occurrences of hexachlorobenzene residues in butter from Spain (n=36, 89% 

positive) and the rest of Europe (n=20, 70% positive) were 5.864±3.171 ng/g wet weight and 

3.022±3.964 ng/g wet weight, respectively (Badia-Vila et al. 2000). 
HEXACHLOROBENZENE 231 

6. POTENTIAL FOR HUMAN EXPOSURE 

The frequency of detection of hexachlorobenzene in the FDA Total Diet Study conducted in 1982–1984 

was 9% (Gunderson 1988). Hexachlorobenzene intakes, in μg/kg body weight/day, estimated for these 

total diet analyses were 0.0020 and 0.0011 for 14–16-year-old males and 60–65-year-old females, 

respectively. In more recent FDA Total Diet Studies, the frequency of detection of hexachlorobenzene 

residues declined to 7% in 1988 (FDA 1989), 5% in 1989 (FDA 1990), 4% in 1990 (FDA 1991), 2% in 

1991 (FDA 1992), <2% on 1991–1993 (FDA 1994), and <2% in 1994 (FDA 1995). Hexachlorobenzene 

intakes (μg/kg body weight/day) estimated for the Total Diet Analyses also declined from intakes 

estimated in the 1982–1984 analysis and were 0.0011 and 0.0006 in 1988 (FDA 1989); 0.0009 and 

0.0005 in 1989 (FDA 1990); 0.0005 and 0.0002 in 1990 (FDA 1991); and 0.0004 and 0.0002 in 1991 

(FDA 1992) for 14–16-year-old males and 60–65-year-old women, respectively. 

Domestic samples of mixed feed rations were collected and analyzed by the FDA for pesticide 

surveillance during fiscal years 1989–1994. Hexachlorobenzene residue was detected in 1 of 457 samples 

in trace amounts (Lovell et al. 1996). 

Burton and Bennett (1987) estimated a human body burden for hexachlorobenzene of 0.7 mg derived 

primarily from dietary intake of fatty foods (0.2 μg/day). Inhalation was estimated to contribute 

100 times less than dietary intake (0.002 μg/day) and consumption of drinking water was also considered 

to contribute only negligible amounts of hexachlorobenzene (0.06 μg/year). 

An exploratory study of chemical exposure was conducted among Vietnamese, Bangladeshi, and local 

resident sportfish consumers in the Montreal region of the St. Lawrence River. The concentration ranges 

for the respective groups are as follows: 0.01–0.04; 0.01–0.02; and 0.01–0.07 μg/L, indicating a positive 

correlation between local residents consuming sportfish from the St. Lawrence River and hexachlorobenzene concentrations (Kosatsky et al. 1999). Anderson et al. (1998) conducted a study to assess hexachlorobenzene contamination in human serum and urine samples from frequent consumers of sport fish 

from Lakes Michigan, Erie, and Huron. A telephone survey was conducted requesting fish eating habits 

with special attention to lake trout, brown trout, rainbow trout, or chinook or coho salmon, carp or catfish, 

and walleye or perch or smelt. After the survey, each angler was invited to give a serum sample. The 

minimum and maximum hexachlorobenzene concentrations of all 30 participating subjects were 0.02 and 

0.2 ppb, respectively, with a median concentration of 0.1 ppb. Eight participants from Lake Michigan 

had minimum and maximum concentrations of 0.09 and 0.2 ppb, respectively, with a median 

concentrations of 0.1 ppb, and 11 participants from Lake Huron and Lake Erie had respective minimum 

and maximum concentrations of 0.04 and 0.2 ppb and 0.02 and 0.2 ppb with median concentrations of 
HEXACHLOROBENZENE 232 

6. POTENTIAL FOR HUMAN EXPOSURE 

0.1 and 0.09 ppb. A comparison group from Arkansas (180 serum samples) had hexachlorobenzene 

concentrations ranging from not detected to 0.3 ppb with a median of 0.1 ppb (Anderson et al. 1998). 

This study illustrates that since the comparison group had the highest range in concentration, there is a 

wide spread of hexachlorobenzene contamination and any population may possibly be affected. 

Sinkkonen et al. (1995b) discovered concentrations of hexachlorobenzene in pine needles in the vicinity 

of a metal reclamation plant. Five sites sampled and analyzed for hexachlorobenzene in the wax of the 

needles and the rest of the needles in 1991 (0.257–0.731 ng/g; 0.758–3.170 ng/g), 1992 

(0.142–0.692 ng/g; 0.255–1.785 ng/g), and 1993 (0 concentration found in the wax of the needles; 

0.217–0.885 ng/g) show decreasing concentrations (Sinkkonen et al. 1995b). In 1996, Sinkkonen et al. 

(1996) analyzed pine needles in and around the metal reclamation plant again. Composite samples from 

the years 1993, 1994, and 1995 had hexachlorobenzene concentrations of 6.9–8.3 ng/g in the wax of the 

needles and 0.2–2.2 ng/g in the rest of the needles, contrary to the decreasing trends found from the 

analysis conducted during 1991–1993. 

6.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSURE 

Hexachlorobenzene is no longer produced (as an end-product) or used as a pesticide in the United States. 

Consequently, the current potential for exposure of the general population appears to be very limited. 

Members of the general population may be exposed to very low concentrations of hexachlorobenzene 

through ingestion of contaminated foodstuffs, particularly those with high lipid content such as meat, 

poultry, and fish. General population exposure to hexachlorobenzene via inhalation or dermal contact 

would be much less. Occupational exposure is possible for workers involved in the production of 

chlorinated hydrocarbons, which releases hexachlorobenzene as a by-product. One recent study has also 

shown farmers to be susceptible to hexachlorobenzene. 

Brock et al. (1998) investigated four families in Iowa and two families in North Carolina for exposure to 

several organochlorine pesticides by analyzing the pesticide levels in their serum. Among the farmers 

from Iowa, mean hexachlorobenzene concentrations in serum ranged from 0.12 to 0.26 ng/mL, and their 

spouses had mean levels ranging from 0.05 to 0.24 ng/mL. The two farmers from North Carolina had 

mean levels of 0.15 and <0.05 ng/mL, and their respective spouses had levels of 0.16 and 0.17 ng/mL. It 

was shown that in one family, the pesticide serum level in the spouse (0.17 ng/mL) was higher than that 

in the farmer (<0.05). Glynn et al. (2000) studied the serum concentration of hexachlorobenzene in a 

group of 790 men ages 40–75 who were randomly selected from both rural and urban areas of Uppsala, 
HEXACHLOROBENZENE 233 

6. POTENTIAL FOR HUMAN EXPOSURE 

Sweden. The mean concentration of hexachlorobenzene was 83.1±133.6 ng/g lipid for this group. This 

study group had lower serum concentrations than males who had recent occupational exposure or high 

environmental exposure. Hagmar et al. (2001) examined blood samples from 110 men who consumed 

varying amounts of fish from the Baltic sea. The median plasma levels of hexachlorobenzene in the 

43 Swedish and 67 Latvian adult males in this study group was 84 ng/g lipid. Recently, the serum 

concentrations of pregnant women from the Disko Bay area, Greenland were studied by Bjerregaard and 

Hansen (2000). The women in this study consume a high level of meat and blubber from marine animals. 

The concentration of hexachlorobenzene in plasma taken from these women between the years 1994 and 

1996 was 1.2 ng/mL wet weight (range, 0.1–7.0 ng/mL wet weight). Hexachlorobenzene concentrations 

in serum in study groups of active smoking mothers, passive smoking mothers, and nonsmoking mothers 

from Germany were 0.87 ng/mL (range, 0.23–4.38 ng/mL), 0.55 ng/mL (range, <0.10–3.27 ng/mL), and 

0.46 ng/mL (range, <0.10–2.73 ng/mL) (Lackman et al. 2000). The mean concentration of hexachlorobenzene in whole blood from infant children ranged from 0.13 to 0.23 ng/mL (Karmaus et al. 2001). The 

highest concentration was observed in children who were breastfed for over 12 weeks after birth. 

Nonoccupational exposure to hexachlorobenzene for residents of two U.S. cities (Jacksonville, Florida 

and Springfield, Massachusetts) were studied over three seasons; summer 1986, spring 1987, and winter 

1988 (Whitmore et al. 1994). The study focused primarily on inhalation exposures with primary 

environmental monitoring consisting of 24-hour indoor, personal, and outdoor air. For the Jacksonville, 

Florida population, the estimated mean hexachlorobenzene concentration ranged from 0.3 to 1.3 ng/m3

 for 

indoor air, from not detected to 0.2 ng/m3

 for outdoor air, and from 0.4 to 0.9 ng/m3 for personal air. For 

the Springfield, Massachusetts population, mean exposures were much less. The estimated hexachlorobenzene concentrations ranged from not detected to 0.1 ng/m3

 for indoor air, were not detected in outdoor 

air, and ranged from not detected to <0.05 ng/m3

 for personal air. The mean air exposure estimated for 

hexachlorobenzene in Jacksonville, Florida, was 10 ng/day, while dietary exposure ranged from 70 to 

120 ng/day. The mean air exposure for Springfield, Massachusetts, was not detected in the personal air 

samplers, while the dietary exposure was 105 ng/day. In both the Jacksonville, Florida population, 

characterized as a high pesticide use area, and in the Springfield, Massachusetts population, characterized 

as a low pesticide use area, the dietary exposure to hexachlorobenzene was the predominant exposure 

pathway. 

In the most recent National Human Adipose Tissue Survey, hexachlorobenzene was found in 35 of 

46 human adipose tissue samples from all regions of the United States at levels ranging from 12 to 

1,300 ppb (EPA 1986f). In other studies of the general population, hexachlorobenzene has been found in 
HEXACHLOROBENZENE 234 

6. POTENTIAL FOR HUMAN EXPOSURE 

human fat samples from residents of the Texas Gulf Coast at concentrations ranging from 18 to 35 ppb 

(Ansari et al. 1986). Kutz et al. (1991) summarized data on hexachlorobenzene residues in human 

adipose tissue collected in the United States from 1973 through 1983. The geometric mean concentrations increased slightly from 0.02 ppm (20 ppb) in 1973 to 0.05 ppm (50 ppb) in 1976, and then declined 

to 0.031 ppm (31 ppb) in 1983. 

Human breast adipose tissue samples from 36 females of 50–80 years in age were collected from the 

Yale-New Haven hospital in Connecticut. A correlation was made between breast adipose tissue and 

serum residues using Pearson’s correlation coefficient. On a lipid adjusted basis, all 36 human adipose 

tissue samples were found to contain residues of hexachlorobenzene. The range of concentration was 

2.5–33.3 ng/g with a median of 17.7 ng/g (Archibeque-Engle et al. 1997). A study conducted in British 

Columbia, Canada, Mes (1992) reported median and maximum hexachlorobenzene residues in biopsied 

fatty tissue of 18.8 and 87 ng/g (ppb), respectively. In a more recent study, small amounts of breast tissue 

were collected from 60 women undergoing breast surgery at Stanford University, California. The mean 

hexachlorobenzene concentration was 46 ng/g fat with a minimum and maximum of 14 and 170 ng/g fat, 

respectively (Petreas et al. 1998). Weistrand and Noren (1998) collected adipose tissue and liver samples 

from five Swedish men and two Swedish women. Hexachlorobenzene concentrations ranged from 12 to 

129 ng/g lipids with a mean of 56 ng/g lipids in adipose tissue and 17 to 156 ng/g lipids with a mean of 

58 ng/g lipids in the liver. Hexachlorobenzene levels in human adipose tissue from 64 mothers living in 

Veracruz, Mexico averaged 0.065 mg/kg (range, 0.010–0.401 mg/kg) on a lipid adjusted basis 

(Waliszewski et al. 2000a). The mean concentrations of hexachlorobenzene in autopsy tissue samples 

from Greenlanders were 594 μg/kg lipid (range, 476–742 μg/kg lipid), 588 μg/kg lipid (range, 

156–1,890 μg/kg lipid), 260 μg/kg lipid (range, 175–387 μg/kg lipid), and 754 μg/kg lipid (range, 

603–943 μg/kg lipid) for subcutaneous fat, omental fat, brain, and liver tissues, respectively (Dewailly et 

al. 1999). 

In the National Health and Nutrition Examination Survey (NHANES II) conducted by the EPA, hexachlorobenzene levels in blood from the general population collected from 1976 to 1980 revealed a median 

concentration of 1.7 ppb, which did not vary among the three age groups studied (Murphy and Harvey 

1993). Hexachlorobenzene levels in normal human blood serum samples were reported to be 2.2 ppb and 

were higher (4.6 ppb) in uremic serum samples (Rutten et al. 1988). In a recent study conducted in 

British Columbia, Canada, Mes (1992), reported median and maximum whole blood levels of 0.11 and 

0.34 ng/g (ppb) in individuals from the general population. 
HEXACHLOROBENZENE 235 

6. POTENTIAL FOR HUMAN EXPOSURE 

Although all uses of hexachlorobenzene as a pesticide in the United States were voluntarily canceled in 

1984, occupational exposures may still occur among workers in the chlorinated solvent manufacturing 

industry, and workers currently involved in the manufacture and application of pesticides contaminated 

with hexachlorobenzene. Military or firefighting personnel who use pyrotechnic mixtures that release 

hexachlorobenzene and workers involved in the disposal of hexachlorobenzene contaminated materials, 

via combustion processes associated with municipal incinerators or those involved in the handling and 

treatment of wastes at hazardous waste sites, may be exposed to higher than background concentrations. 

No information was located on concentrations of hexachlorobenzene in occupationally exposed 

populations in the United States. However, in a 10-year study (1976–1985), human adipose tissue 

samples and human milk from patients exposed to PCBs or pesticides in Ontario, Canada were analyzed 

for a variety of pesticides and industrial chemicals (Frank et al. 1988). Residues of hexachlorobenzene in 

adipose tissues averaged at or below 0.2 ppb in the extractable fat. Concentrations of hexachlorobenzene 

in milk ranged from a mean of 0.52 ppb (in whole milk) in 1983–1984 to a mean of 0.26 ppb (in whole 

milk) in 1985. The highest mean concentration (1.33 ppb) was observed in central Ontario residents. 

Urban residents had a higher mean concentration (0.57 ppb) as compared to rural residents with a mean of 

0.27 ppb. 

Plasma hexachlorobenzene concentrations in a Louisiana population living in a hexachlorobenzenecontaminated area averaged 3.6±4.3 ppb. The highest plasma level (345 ppb) was detected in a waste 

disposal facility worker, while the highest plasma level in the general population was 23 ppb (Burns and 

Miller 1975). Workers at a carbon tetrachloride and perchloroethylene production facility had plasma 

hexachlorobenzene levels of up to 223 ppb. Hexachlorobenzene blood levels were determined over a 

4-year period in men employed in the manufacture of chlorinated solvents (Currier et al. 1980). Blood 

levels ranged from 5 to 1,121 ppb (310.7 ppb mean) in 1974, 30–990 ppb (311.5 mean) in 1975, 

3–600 ppb (159.9 mean) in 1976, and 22–467 ppb (170.3 mean) in 1977. The hexachlorobenzene blood 

levels were strongly associated with the number of years worked in the chlorinated solvents plant, but 

they were poorly correlated with airborne hexachlorobenzene concentrations ranging from <1 to 13 ppb 

or wipe samples from work areas ranging from 0.03 to 124 μg/100 m2

. 

Vegetable sprayers applying hexachlorobenzene-contaminated dimethyltetrachloroterephthalate (DCPA) 

had plasma levels of hexachlorobenzene ranging from 0 to 310 ppb (mean 40±63 ppb), accompanied by 

elevated levels of delta-aminolevulinic acid, but no health related adverse effects (Burns et al. 1974). 
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6. POTENTIAL FOR HUMAN EXPOSURE 

Elevated urinary uroporphyrin and coproporphyrin were found in 1 of 54 men occupationally exposed to 

hexachlorobenzene (Morley et al. 1973). 

Workers at a new hazardous waste incinerator in Constanti, Spain had mean plasma levels of hexachlorobenzene at 152 μg/kg lipid (range, 19.4–854.0 μg/kg lipid) (Domingo et al. 2001). Residents (n=608) 

living near an electrochemical factory in Catalonia, Spain had mean serum concentrations of hexachlorobenzene as follows (units ng/mL): general population, male (50.2), female (48.0), all (48.9); nonfactory 

workers, male (39.8), female (47.9), all (46.3); and living with a worker of the factory, yes (46.8), no 

(46.8) (Ballester et al. 2000). 

Individuals employed in industries that manufacture or process hexachlorobenzene or products containing 

hexachlorobenzene may be exposed to the highest concentrations. The National Occupational Exposure 

Survey (NOES), conducted by NIOSH from 1981 to 1983, estimated that 1,038 workers employed at 

10 facilities were potentially exposed to hexachlorobenzene in the United States (NOES 1990). The 

NOES database does not contain information on the frequency, concentration, or duration of exposure; 

the survey provides only estimates of workers potentially exposed to chemicals in the workplace. 

6.6 EXPOSURES OF CHILDREN 

This section focuses on exposures from conception to maturity at 18 years in humans. Differences from 

adults in susceptibility to hazardous substances are discussed in 3.7 Children’s Susceptibility. 

Children are not small adults. A child’s exposure may differ from an adult’s exposure in many ways. 

Children drink more fluids, eat more food, breathe more air per kilogram of body weight, and have a 

larger skin surface in proportion to their body volume. A child’s diet often differs from that of adults. 

The developing human’s source of nutrition changes with age: from placental nourishment to breast milk 

or formula to the diet of older children who eat more of certain types of foods than adults. A child’s 

behavior and lifestyle also influence exposure. Children crawl on the floor, put things in their mouths, 

sometimes eat inappropriate things (such as dirt or paint chips), and spend more time outdoors. Children 

also are closer to the ground, and they do not use the judgment of adults to avoid hazards (NRC 1993). 

Children are particularly susceptible to hexachlorobenzene by ingestion of breast milk. Table 6-5 

summarizes the concentrations of hexachlorobenzene in human breast milk taken from women living in 

different regions of the world. Hexachlorobenzene was detected in human milk samples collected from 
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6. POTENTIAL FOR HUMAN EXPOSURE 

54 residents of Hawaii during 1979–1980 (Takei et al. 1983). The incidence of detection of hexachlorobenzene in the sampled population was 100% and the mean concentration of positive detections was 

46±49 ppb (ng/g lipid basis) with residues ranging from 18 to 38 ppb (ng/g lipid basis). The authors state 

that the levels of hexachlorobenzene in human milk from residents of Hawaii are consistent with levels 

detected in an earlier human milk study conducted on women on the mainland United States. Schecter et 

al. (1998) found that hexachlorobenzene residues in the breast milk of a mother with nursing twins 

decreased from 10.7 ng/g lipid to not detectable in 30 months. Thus, nursing infants are particularly 

susceptible to hexachlorobenzene poisoning due to the mother’s decrease in body burden and the infant’s 

intake (Schecter et al. 1998). A recent study of organochlorine pesticide concentrations in human milk 

sampled throughout Canada during 1992 found that hexachlorobenzene was present in all 497 milk 

samples at a mean concentrations of 0.44 ng/g (ppb) in whole milk and 14.5 ng/g (ppb) in milk fat 

(Newsome et al. 1995). A comparison by Canada Health of human milk contamination in whole milk 

from 1967 to 1992 was conducted in Canada, Quebec and Ontario. This study showed a decrease in 

hexachlorobenzene concentration in Canada (mean concentration; 2–0.44 ng/g), Quebec (median 

concentration; 2–0.40 ng/g), and Ontario (median concentration; 2–0.48 ng/g) (Craan and Haines 1998). 

Newsome et al. (1995) reported that concentrations of hexachlorobenzene were higher in women from 
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6. POTENTIAL FOR HUMAN EXPOSURE 

Table 6-5. Mean Levels of Hexachlorobenzene in Breast Milk 

Location of study N/n Concentrationa Reference 

New Zealand Bates et al. (1994) 

Auckland, urban 11/n/a 0.020 

Northland, rural 10/n/a 0.021 

Christchurch, urban 9/n/a 0.030 

Canterbury, rural 8/n/a 0.063 

Porto Alegre, Brazil 30/19 0.02 Beretta and Dick (1994) 

France (multiple locations) 20/19 0.147 Bordet et al. (1993) 

Madrid, Spain 63/52 1.0 Conde et al. (1993) 

Industrialized area 1.74 

Canada (1967–1992) no data 0.002–0.00044b Craan and Haines (1998) 

Quebec 0.002–0.00040b 

Ontario 0.002–0.00048b 

Arctic Quebec Dewailly et al. (1993) 

Inuit women 107/107 0.136 

Caucasian women 50/48 0.028 

Italy 27/no data Larsen et al. (1994) 

Rome 0.25 

Pavia 0.12 

Milan 0.20 

Florence 0.14 

Arkansas, USA 942/57 0.03 Mattison et al. (1992) 

Canada 412/no data 0.026 Mes et al. (1993) 

Victoria, Australia (1985/1986) 158/153 0.005b Monheit and Luke (1990) 

Canada (multiple locations, in 497/497 0.0044b Newsome et al (1995) 

1992) 0.015c
 

Sweden (1972–80) 227–1,500/ 0.110–220 Norén and Meironyté (2000)
 

no data 

Akumadan, Ghana 20/19 0.04 Ntow (2001) 

Bratislava, Slovak Republic 26/26 0.339 Prachar et al (1993) 

Victoria, Australia 60/59 0.41 Quinsey et al (1995) 

Czech Republic Schoula et al (1996) 

Prague 17 0.639
 

Kladno 12 0.570
 

Uherske Hradiste 7 0.482
 

Northern Thailand 25/9 0.0051b Stuetz et al (2001) 

Hawaii, USA (1979–1980) 54/54 0.046±0.049 Takei et al. (1983) 

Veracruz, Mexico 43/43 0.047 Waliszewski et al (1996) 

Source: Pohl and Tylenda (2000) 

a

μg/g on lipid basis

b

whole milk sample 

c

milk fat sample 

N = number of samples; n = number of samples with measurable levels 
HEXACHLOROBENZENE 239 

6. POTENTIAL FOR HUMAN EXPOSURE 

the Great Lakes Basin area as compared to the rest of Canada. Concentrations of hexachlorobenzene 

were higher in the breast milk of women who consumed more than 100 g of fish weekly. 

Children may also be exposed to chemicals via ingestion of contaminated foods. Hexachlorobenzene 

residues have been detected in 76% of samples analyzed as part of the National Human Adipose Tissue 

Survey (FY82) (EPA 1986f). These hexachlorobenzene residues are most likely the result of 

consumption of low levels of hexachlorobenzene in food, with calculated yearly intakes of 68, 22, and 

5 μg for adults, toddlers, and infants, respectively (EPA 1986e). More recently, Yess et al. (1993) 

evaluated hexachlorobenzene residues from 1985 to 1991 detected in the Total Diet Studies of infant and 

adult foods that are consumed by infants and young children. These authors reported maximum hexachlorobenzene residues detected in various food groups as follows: combination meat dinners—pork 

(0.4 ppb), beef (0.3 ppb), chicken/turkey (0.3 ppb), chicken/turkey/vegetable (0.3 ppb), beef and 

vegetable (0.1 ppb); vegetables and fruits—pears (1.0 ppb), apples (0.4 ppb), and carrots (0.2 ppb); milk 

products—canned evaporated milk (0.5 ppb), whole milk (0.2 ppb), and low-fat (2%) milk (0.1 ppb); and 

peanut butter (5.0 ppb). 

Hexachlorobenzene intakes, in μg/kg body weight/day, estimated for these total diet analyses 

(1982–1984) were 0.0015 for 6–11-month-old infants. A follow-up study found a decrease in intakes that 

were estimated in 1982–1984. Hexachlorobenzene intakes(μg/kg body weight/day) were estimated to be 

0.0016 in 1988 (FDA 1989); 0.0007 in 1989 (FDA 1990); 0.0004 in 1990 (FDA 1991); and 0.0003 in 

1991 (FDA 1992) for 6–11-month-old infants. 

Although inhalation exposures of hexachlorobenzene in children have not been studied, it is anticipated 

that exposure by this route will not be significant in outdoor environments. The Henry’s law constant of 

hexachlorobenzene is 5.8x10-4 atm-m3

/mol (Ten Hulscher et al. 1992), indicating that this compound will 

volatilize rapidly, especially in moist soils with low organic content. Hexachlorobenzene’s high log Koc 

of 6.08 (EPA 1981a), however, indicates that volatilization from soil surfaces will be greatly attenuated. 

Considering that hexachlorobenzene concentrations in the environment are extremely low, exposure of 

children by inhalation is expected to be insignificant. After a hexachlorobenzene spill, inhalation 

exposure may be important before environmental equilibrium is attained. Under these conditions, high 

concentrations of hexachlorobenzene would be found in the atmosphere, due to hexachlorobenzene’s 

calculated vapor density of 10. This situation, however, is not expected to occur since hexachlorobenzene 

is no longer produced or used commercially and is only found as an impurity in pesticides and as a byproduct of chlorinated hydrocarbons. 
HEXACHLOROBENZENE 240 

6. POTENTIAL FOR HUMAN EXPOSURE 

The EPA issued a warning regarding pesticides and advised that potential exposure of pesticides to young 

children via dermal absorption and ingestion was more important than inhalation routes (Jantunen et al. 

1997). 

6.7 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES 

In addition to those individuals occupationally exposed to hexachlorobenzene identified in Section 6.5, 

several groups within the general population may receive potentially higher exposures to hexachlorobenzene. These groups within the general population include individuals living near facilities where 

hexachlorobenzene is produced as a byproduct, individuals living near the 101 current or former NPL 

hazardous waste sites where this compound is present, recreational and subsistence fishermen who 

consume higher amounts of fish than the general population, and native populations (including Native 

American populations such as the Inuits of Alaska) who may be exposed to higher levels of hexachlorobenzene associated with dietary intakes of caribou and other game species. 

Consumption of contaminated groundwater by individuals living in the vicinity of production facilities or 

hazardous waste sites may be an important source of exposure for both adults and children. Hexachlorobenzene has been detected in 42 groundwater samples from 107 NPL hazardous waste sites (HazDat 

2002). Skin contact with or ingestion of hexachlorobenzene contaminated soil on their hands may be an 

additional source of exposure for children living near hazardous waste facilities. Hexachlorobenzene has 

been detected in 76 soil samples from 107 NPL hazardous waste sites (HazDat 2002). 

Recreational and subsistence fishermen who consume appreciably higher amounts of locally caught fish 

from contaminated water bodies may be exposed to higher levels of hexachlorobenzene associated with 

dietary intake than members of the general population (EPA 1995b). Hexachlorobenzene contamination 

in fish and shellfish has triggered the issuance of several human health advisories. As of September 1994, 

hexachlorobenzene was identified as the causative pollutant in fish consumption advisories in Louisiana 

and Ohio. This information is summarized in Table 6-6. EPA has identified hexachlorobenzene as a 

target analyte and recommended that this chemical be monitored in fish and shellfish tissue samples 

collected as part of state toxics monitoring programs. EPA recommends that 
HEXACHLOROBENZENE 241 

6. POTENTIAL FOR HUMAN EXPOSURE 

Table 6-6. Fish Consumption Advisoriesa 

State Waterbody Extent Species 

Louisiana Devil’s Swamp Lake and Bayou 

Baton Rouge 

7 miles All fish 

Calcasieu and Cameron 

parishes 

6 miles All fish and shellfish 

Ohio Tuscarawas River Turkey Foot Rd. (SR 

619), Barberton, to South 

Broadway St (SR 416), 

New Philadelphia 

(Tuscarawas County) 

Carp-common, basssmallmouth, bullheadyellow, bass-rock, 

catfish-channel, basslargemouth 

a

From EPA 1997b National Listing of Fish Consumption Advisories 
HEXACHLOROBENZENE 242 

6. POTENTIAL FOR HUMAN EXPOSURE 

residue data obtained from these monitoring programs should then be used by the states to conduct risk 

assessments to determine the need for issuing fish and shellfish consumption advisories for the protection 

of the general public as well as recreational and subsistence fishermen (EPA 1997b). 

Native American populations such as the Inuits of Alaska or other subsistence hunters living in high 

latitude areas of the United States and Canada may be exposed to hexachlorobenzene residues in caribou, 

beluga whales, polar bears, seals, and other game species. Significantly higher concentrations of hexachlorobenzene (mean 136 ng/g [ppb]) were reported in breast milk of Inuit mothers from eastern Canada 

(Quebec Province) as compared with residues of 28 ng/g (ppb) in Caucasian mothers (Dewailly et al. 

1993). By analogy, it is possible that Inuit populations in western North America (Alaska) may receive 

potentially higher hexachlorobenzene exposures from their dietary habits. In a follow-up study by 

Dewailly et al. (1999) 26 subcutaneous fat samples, 41 omental fat samples, 17 brain samples, and 

26 liver samples were collected in November 1992 to Mid-October 1994 from Inuit Greenlanders. Mean 

hexachlorobenzene concentrations were 594, 588, 260, and 754 μg/kg lipid basis, respectively. A 

comparison of these data clearly suggest an increase in Inuit population’s hexachlorobenzene levels from 

dietary habits. Maternal body burden and lactational transfer of hexachlorobenzene can increase tissue 

levels in the neonate (Ando et al. 1985; Frank et al. 1988). 

6.8 ADEQUACY OF THE DATABASE 

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with the 

Administrator of EPA and agencies and programs of the Public Health Service) to assess whether 

adequate information on the health effects of hexachlorobenzene is available. Where adequate 

information is not available, ATSDR, in conjunction with the National Toxicology Program (NTP), is 

required to assure the initiation of a program of research designed to determine the health effects (and 

techniques for developing methods to determine such health effects) of hexachlorobenzene. 

The following categories of possible data needs have been identified by a joint team of scientists from 

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would 

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean 

that all data needs discussed in this section must be filled. In the future, the identified data needs will be 

evaluated and prioritized, and a substance-specific research agenda will be proposed. 
HEXACHLOROBENZENE 243 

6. POTENTIAL FOR HUMAN EXPOSURE 

6.8.1 Identification of Data Needs 

Physical and Chemical Properties. The physical and chemical properties of hexachlorobenzene 

are sufficiently well documented to permit estimation of the compound's environmental fate (Budavari 

1996; EPA 1981a; Hansch et al. 1995; Lide 1998; Ten Hulscher et al. 1992; Verschueren 1996). No 

further information is needed. 

Production, Import/Export, Use, Release, and Disposal. Hexachlorobenzene is not currently 

manufactured as a commercial end-product in the United States and has not been commercially produced 

since the late 1970s (EPA 1986e). However, hexachlorobenzene currently is produced as a by-product or 

impurity in the manufacture of chlorinated solvents, other chlorinated compounds, and several currently 

registered pesticides (Bailey 2001; EPA 1986e, 1994c; IARC 1979). 

The total amount of hexachlorobenzene released as a by-product in the production of all chlorinated 

solvents was estimated as 0.3 kg/year in the mid 1990s (Bailey 2001), while hexachlorobenzene released 

through use of eight major pesticides containing hexachlorobenzene accounted for 1,270 kg/year (Bailey 

2001). Current production estimates for hexachlorobenzene as a by-product or impurity are not available. 

Current, quantitative estimates of production of hexachlorobenzene from all sources are needed to 

evaluate potential exposures and risks to human health. 

There are no current commercial uses of hexachlorobenzene in the United States, although the compound 

was used as a fungicide until 1984 when the last registered use as a pesticide was voluntarily cancelled 

(Beyer 1996). Prior to the registration cancellations, hexachlorobenzene was registered as a seed 

protecting for use on several grains (principally wheat) and field crops (EPA 1986e). Hexachlorobenzene 

was also used in pyrotechnic and ordinance materials and in synthetic rubber production (EPA 1986e). 

Impurities of hexachlorobenzene in currently registered pesticides (picloram, PNCB, chlorothalonil, 

Dacthal®, atrazine, simazine, lindane, and PCP) (Bailey 2001; EPA 1986e, 1993a; Farm Chemicals 

Handbook 1993) appear to be a continuing source of hexachlorobenzene exposure for the general 

population. Five of the pesticides containing impurities of hexachlorobenzene (PCNB, chlorothalonil, 

Dacthal®, lindane, and PCP) are used in home gardens, lawn care, and other applications around 

residences and in urban areas (Farm Chemicals Handbook 1993). Currently, data regarding the release of 

hexachlorobenzene into the environment as an impurity via the use of other pesticides is not available. 
HEXACHLOROBENZENE 244 

6. POTENTIAL FOR HUMAN EXPOSURE 

No current information was located on import/export volumes for hexachlorobenzene, although
 

import/export volumes for hexachlorobenzene/DDT combined are available (NTDB 1995).
 

Hexachlorobenzene is listed as a hazardous waste. It is regulated under the Clean Water Effluent
 

Guidelines as stated in Title 40, Section 400–475, of the Code of Federal Regulations and the Resource
 

Conservation and Recovery Act (RCRA) (see Chapter 8). Past disposal methods have included
 

incineration, landfills, discharges to municipal sewage treatment plants, and emissions to the atmosphere. 

The recommended method of disposal for hexachlorobenzene is incineration (Clayton and Clayton 1981;
 

EPA 1988a, 1989a; IRPTC 1985; Lamb et al. 1994). No further information on disposal practices is
 

needed; however, estimates on the volume of hexachlorobenzene disposed of annually and the disposal
 

method used are needed to assess exposure pathways. 

According to the Emergency Planning and Community Right-to-Know Act of 1986, 42 U.S.C.
 

Section 11023, industries are required to submit chemical release and off-site transfer information to the
 

EPA. The TRI, which contains this information for 1999, became available in 2001. This database will
 

be updated yearly and should provide a list of industrial production facilities and emissions.
 

Environmental Fate. Hexachlorobenzene released to the environment partitions to several
 

environmental compartments (air, water, soil and sediment, and biological organisms). Hexachlorobenzene partitions to the atmosphere from soil surfaces through volatilization (Nash and Gish 1989). The
 

remainder is adsorbed strongly to soil where it persists for extended periods (half-life of months to years)
 

due to its resistance to biodegradation (Beall 1976; Beck and Hansen 1974; Isensee et al. 1976). 

Leaching of hexachlorobenzene into groundwater is not expected to occur very rapidly under most
 

circumstances due to the compound's high sorption characteristics (Swann et al. 1983). Yuan et al. (1999)
 

have recently reported that hexachlorobenzene can be dechlorinated to dichlorobenzenes under anaerobic
 

conditions in the laboratory using sewage sludge as inoculum.
 

Once in the atmosphere, hexachlorobenzene exists in both the vapor and particulate phase; however, the
 

vapor phase predominates (Ballschmiter and Wittlinger 1991; Bidleman et al. 1989; Lane et al. 1992). 

Degradation of hexachlorobenzene in the atmosphere is quite slow (1.69 years) (Brubaker and Hites
 

1998). Since hexachlorobenzene is hydrophobic, wet deposition will not be an important loss process. In
 

cold high latitude zones, dry deposition of hexachlorobenzene aerosols is encouraged (Ballschmiter and
 

Wittlinger 1991; Lane et al. 1992; Wania and Mackay 1993). Atmospheric transport of hexachlorobenzene is a major mechanism for global translocation of this compound (Eisenreich et al. 1981; Kelly et
 
HEXACHLOROBENZENE 245 

6. POTENTIAL FOR HUMAN EXPOSURE 

al. 1991). Long-range global transport of hexachlorobenzene released anywhere in the world can occur 

via atmospheric or oceanic systems (Ballschmiter and Wittlinger 1991; Wania and MacKay 1993). 

Hexachlorobenzene released to water will volatilize, adsorb to sediments, or bioaccumulate in fish and 

other aquatic organisms (Bishop et al. 1995; EPA 1992b; Kelly et al. 1991; Langlois and Langis 1995; 

Oliver and Nichol 1982a; Quemerais et al. 1994; Rostad et al. 1993; Schmitt et al. 1990; Zabik et al. 

1995). Hydrolysis and biodegradation are not significant processes in water. Information on biodegradation of hexachlorobenzene under anaerobic conditions in a laboratory study exists (Yuan et al. 1999), 

but degradation under field conditions were not found. Further information on these processes, including 

degradation products, are needed to determine potential mechanisms and sources of hexachlorobenzene 

releases from soils and the potential for the compound and its degradation products to contaminate 

groundwater. 

Both bioaccumulation and biomagnification of hexachlorobenzene were reported to occur in an aquatic 

laboratory microcosm system (Burkhard et al. 1997; Isensee et al. 1976); however, data by Russell et al. 

(1995) suggests that hexachlorobenzene bioaccumulates, but is not biomagnified in certain fish 

populations in Lake Erie. In terrestrial ecosystems, hexachlorobenzene can also be accumulated in 

several agricultural plant species in the roots and parts of the plants closest to the soil (Kraaij and Connell 

1997; Scheunert et al. 1983; Schroll et al. 1994; Smelt and Leistra 1974). In lichens, a high latitude 

forage food for caribou, hexachlorobenzene was found to be bioconcentrated 8,800,000–17,000,000 times 

the concentration in the atmosphere (Muir et al. 1993). Although the issue of biomagnification in some 

ecosystems needs to be clarified, there are adequate data on the bioconcentration of hexachlorobenzene in 

both aquatic and terrestrial ecosystems. 

Bioavailability from Environmental Media. Hexachlorobenzene can be absorbed following 

inhalation of contaminated workplace air (Burns et al. 1974; Currier et al. 1980; Richter et al. 1994). 

Since hexachlorobenzene is only moderately volatile, inhalation may not be a major concern except at 

hazardous waste sites or in industrial settings. Hexachlorobenzene can be absorbed following ingestion 

of contaminated food or water. Exposure to hexachlorobenzene through ingestion of food contaminated 

with low levels of the compound is probably the greatest source of exposure for the general population. 

Exposure to hexachlorobenzene through ingestion of contaminated drinking water is not expected to be 

an important source of concern since the compound is not very soluble in water. Although there are no 

quantitative data on the human absorption of orally administered hexachlorobenzene, gastrointestinal 

absorption has been demonstrated for rats (Albro and Thomas 1974; Ingebrigtsen and Nafstad 1983; 
HEXACHLOROBENZENE 246 

6. POTENTIAL FOR HUMAN EXPOSURE 

Ingebrigtsen et al. 1981). The lymphatic system has also been shown to play an important part in the 

absorption of hexachlorobenzene in the intestines. Hexachlorobenzene is absorbed by the lymphatic 

system in the region of the duodenum and jejuno-ileum and is deposited in the adipose tissue, bypassing 

the portal circulation (Iatropoulos et al. 1975). Since hexachlorobenzene is tightly bound to soil particles, 

ingestion of hexachlorobenzene-contaminated soil, particularly by children, may also be an important 

route of exposure near production and processing facilities or near hazardous waste disposal sites. No 

information was available regarding absorption of hexachlorobenzene following dermal contact. 

Information regarding the bioavailability of hexachlorobenzene from both ingestion of soil-bound hexachlorobenzene particularly in children and from dermal contact with contaminated soils are needed, 

particularly in assessing health risks to populations living near hazardous waste sites. 

Food Chain Bioaccumulation. Like many of the other organochlorine pesticides, hexachlorobenzene is lipophilic and has a high bioaccumulation potential. Hexachlorobenzene is bioaccumulated in 

fish and other aquatic organisms (Bishop et al. 1995; EPA 1992b; Langlois and Langis 1995; Murray et 

al. 1980, 1981; Oliver and Nichol 1982a; Schmitt et al. 1990; Zabik et al. 1995) as well as waterfowl 

(Cobb et al. 1994; Foley 1992; Gebauer and Weseloh 1993; Somers et al. 1993; Swift et al. 1993; 

Yamashita et al. 1992). Hexachlorobenzene is bioaccumulated in aquatic food chains with virtually no 

degradation of the compound by the exposed organisms (Isensee et al. 1976). The results of a laboratory 

aquatic ecosystem study suggest that bioaccumulation as well as biomagnification of hexachlorobenzene 

occurs (Isensee et al. 1976); however, authors of a more recent study conducted in natural fish 

populations suggest that hexachlorobenzene is not biomagnified (Russell et al. 1995). In terrestrial 

ecosystems, several agricultural crops have been found to accumulate hexachlorobenzene in their roots 

and in portions growing closest to soil level (Ecker and Horak 1994; Scheunert et al. 1983; Schroll et al. 

1994; Smelt and Leistra 1974). The edible root portion of carrots accumulated the highest hexachlorobenzene concentration with a BCF of 19 (Smelt and Leistra 1974). Lichens, a primary forage for caribou, 

were also shown to bioaccumulate hexachlorobenzene (Muir et al. 1993). A field study on a terrestrial 

ecosystem suggested that hexachlorobenzene was biomagnified through various trophic levels of the food 

web (Hebert et al. 1994). Further studies are needed to resolve whether hexachlorobenzene is 

biomagnified in both aquatic and terrestrial ecosystems. 

Exposure Levels in Environmental Media. Environmental monitoring data are available for 

hexachlorobenzene in air (Currier et al. 1980; Davis and Morgan 1986; Eisenreich et al. 1981; EPA 

1975b, 1985g; Hoff et al. 1996; Lee et al. 2000; Mann et al. 1974; Poissant et al. 1997; Tiernan et al. 

1985), water (Chan et al. 1994; Davis and Morgan 1986; EPA 1976a; Hoff et al. 1996; Laska et al. 1976; 
HEXACHLOROBENZENE 247 

6. POTENTIAL FOR HUMAN EXPOSURE 

Oliver and Nichol 1982a; Quemerais et al. 1994), soil (Elder et al. 1981; EPA 1985g; Laseter et al. 1976; 

Laska et al. 1976), and sediment (Davis and Morgan 1986; Elder et al. 1981; Murray et al. 1981; Oliver 

and Nichol 1982a; Ray et al. 1983; Rostad et al. 1999). Groundwater monitoring data are lacking; 

however, hexachlorobenzene has been detected in 42 groundwater samples from 107 NPL hazardous 

waste sites. Current information on hexachlorobenzene concentrations in groundwater is needed. 

Human intake estimates for exposure from environmental media are available (Whitmore et al. 1994), but 

are limited. In general, while monitoring data are available for most environmental media, much of the 

information on environmental levels was collected during the 1970s through the mid 1980s. More recent 

monitoring data from all environmental media would provide more accurate information for estimating 

human and animal intakes. 

Reliable monitoring data for the levels of hexachlorobenzene in contaminated media at hazardous waste 

sites are needed so that the information obtained on levels of hexachlorobenzene in the environment can 

be used in combination with the known body burden of hexachlorobenzene to assess the potential risk of 

adverse health effects in populations living in the vicinity of hazardous waste sites. 

Exposure Levels in Humans. Hexachlorobenzene has been detected in human adipose tissue 

(Ansari et al. 1986; EPA 1986f; Frank et al. 1988; Kutz et al. 1991; Mes 1992;), blood (Burns and Miller 

1975; Burns et al. 1974; Currier et al. 1980; Murphy and Harvey 1993; Rutten et al. 1988), and milk 

(Craan and Haines 1998; Frank et al. 1988; Newsome et al. 1995; Schecter et al. 1998; Takei et al. 1983). 

Studies exist that relate occupational exposure to blood levels of hexachlorobenzene (Burns et al. 1974; 

Currier et al. 1980). Studies to compare the steady-state intake of hexachlorobenzene as measured by 

urinary and fecal excretion as it relates to blood levels in occupationally exposed workers would be 

particularly useful. Since hexachlorobenzene has been detected in both urine and feces, a study of this 

nature could be conducted. These studies might also address possible individual differences in the 

metabolism of this compound. This information is necessary for assessing the need to conduct health 

studies on these populations. 

Exposures of Children. Exposure pathways for children have been well documented in breast milk 

(Newsome et al. 1995; Schecter et al. 1998) and diet (FDA 1992; Yess et al. 1993). Data documenting 

body burdens for children are needed. Data addressing exposure to children who live, play, or attend 

school near NPL sites, industrial sites, such as chlorinated hydrocarbon production factories, or on 

farmlands where hexachlorobenzene is being released as an impurity of another pesticide would allow for 

a better assessment of hexachlorobenzene exposure. As hexachlorobenzene is released due to the use of 
HEXACHLOROBENZENE 248 

6. POTENTIAL FOR HUMAN EXPOSURE 

other pesticides on foods, an evaluation of possible hexachlorobenzene residues in children’s food 

substances would further enhance the ability to construct a complete picture of exposure. Studies 

revealing contamination of drinking water or groundwater would also prove essential in this assessment. 

Attention should also focus on the use of tap water as a contaminant source when used to prepare infant 

formulas from condensed or powdered forms. As children are often bound to pick up soil off the ground 

and maybe even put this soil in their mouths, studies regarding exposure to children through soil would be 

helpful. Information concerning childhood-specific means to decrease exposure would be useful. 

Child health data needs relating to susceptibility are discussed in 3.12.2 Identification of Data Needs: 

Children’s Susceptibility. 

Exposure Registries. No exposure registries for hexachlorobenzene were located. This substance is 

not currently one of the compounds for which a subregistry has been established in the National Exposure 

Registry. The substance will be considered in the future when chemical selection is made for 

subregistries to be established. The information that is amassed in the National Exposure Registry 

facilitates the epidemiological research needed to assess adverse health outcomes that may be related to 

exposure to this substance. Currently, Inuit communities appear to have potentially higher exposures to 

hexachlorobenzene and should be further monitored (Dewailly et al. 1993, 1999). 

6.8.2 Ongoing Studies 

As part of the Third National Health and Nutrition Evaluation Survey, the Environmental Health 

Laboratory Sciences Division of the National Center for Environmental Health and Injury Control, 

Centers for Disease Control and Prevention, will be analyzing human blood samples for hexachlorobenzene and other volatile organic compounds. These data will give a more current indication of the 

frequency of occurrence and background levels of these compounds in the general population. 

According to FEDRIP (2001), there is one ongoing study on hexachlorobenzene being conducted at the 

Virginia Polytechnic Institute. This study involves an effort to reveal the aryl dehalogenation mechanism 

and to evaluate the ability of carbon monoxide reduced carbon monoxide dehydrogenase to reductively 

dehalogenate hexachlorobenzene. 

HEXACHLOROBENZENE 250 

7. ANALYTICAL METHODS 

The purpose of this chapter is to describe the analytical methods that are available for detecting, 

measuring, and/or monitoring hexachlorobenzene, its metabolites, and other biomarkers of exposure and 

effect to hexachlorobenzene. The intent is not to provide an exhaustive list of analytical methods. 

Rather, the intention is to identify well-established methods that are used as the standard methods of 

analysis. Many of the analytical methods used for environmental samples are the methods approved by 

federal agencies and organizations such as EPA and the National Institute for Occupational Safety and 

Health (NIOSH). Other methods presented in this chapter are those that are approved by groups such as 

the Association of Official Analytical Chemists (AOAC) and the American Public Health Association 

(APHA). Additionally, analytical methods are included that modify previously used methods to obtain 

lower detection limits and/or to improve accuracy and precision. 

7.1 BIOLOGICAL MATERIALS 

Methods for the determination of organochlorine compounds such as hexachlorobenzene generally 

consist of the following steps: extraction of the analyte from the sample matrix; clean-up to remove 

interfering compounds; and analysis (separation and quantitation). The primary method of analysis is gas 

chromatography (GC) coupled with electron capture detection (ECD) or mass spectrometry (MS). 

Analytical methods have been developed for the determination of hexachlorobenzene in blood or serum, 

urine, feces, adipose tissue, and breast milk. A summary of methods is shown in Table 7-1. 

Several cautions should be noted. Interferences may result from organics of biological origin that are 

extracted from the sample, and from contaminated glassware, solvent, etc. Sample interferences are 

usually removed using fractionation and clean-up procedures. Rigorous sample collection and 

preparation methods must be followed to prevent contamination of the sample. Good quality control 

procedures must be used to identify and remove interferences caused by sample contamination. 

Blood (or serum) is a body fluid often utilized to assess human exposure to chlorinated organics, 

including hexachlorobenzene. Blood is usually extracted with solvent (Bristol et al. 1982; Burse et al. 

1990; EPA 1980c; Langhorst and Nestrick 1979; Mes et al. 1982), and the extract is cleaned up (and 

sometimes fractionated) by column chromatography utilizing silica gel (Langhorst and Nestrick 1979), 
Table 7-1. Analytical Methods for Determining Hexachlorobenzene in Biological Sa

Sample matrix Preparation method Analytical method 

Sample detection limit 

Percent recovery 

Adipose tissue Extraction, GPC clean-up, 

Florisil fractionation, optional additional clean-up 

cap. GC/MS 12 ng/g No data 

Adipose tissue Maceration with sodium sulfate, 

extraction and back extraction, Florisil fractionation 

GC/ECD No data No data 

Adipose tissue Soxhlet extraction, clean-up 

on Florisil 

cap. GC/ECD; confirmation on second column 

0.001 μg/g 82 

Adipose tissue Solvent extraction, filtration, Florisil 

fractionation 

cap. GC/ECD; confirmation by GC/MS 

0.12 ng/g 86 

Adipose tissue SFE with alumina (to remove lipids, 

purification by column chromatography 

cap. GC/ECD 10 μg/kg 

(fatty tissue) 

115 

Breast milk Separation of fat; column clean-up cap GC/ECD 0.4 ng/g fat No data Breast milk Acid treatment, elute from silica gel, 

concentrate 

GC/ECD 0.009 mg/kg 91 

Blood Solvent (hexane) extraction, concentration GC/ECD No data No data Blood Solvent extraction, clean up on silica gel, 

concentration 

GC/PID 16 ng/g 79 

Blood Homogenization with benzene, filtration, 

Florisil fractionation 

cap. GC/ECD; confirmation by GC/MS 

0.2 ng/g 80 
Table 7-1. Analytical Methods for Determining Hexachlorobenzene in Biological Samples 

Sample detection Percent 

Sample matrix Preparation method Analytical method limit recovery 

Blood Hexane extraction, concentration GC/ECD; 

confirmation by GC/MS 

0.16 ng/g 72 

Serum Solvent extraction of denatured serum, 

fractionation on micro-Florisil column, acid treatment/silica gel clean-up 

GC/ECD 1 ppb 58–76 

Urine Solvent extraction, clean-up on silica gel, 

concentration 

GC/PID 4.1 ng/g 84 

Semen Solvent extraction, clean-up on Florisil, 

concentration 

cap. GC/ECD; confirmation by NICI 

.0.3 ng/mL 80 

Feces Boiling with solvent, clean-up on alumina cap. GC/ECD No data No data 

cap. = capillary; ECD = electron capture detector; GC = gas chromatography; GPC = gel permeation chromatography; MS 

NICI = negative ionization chemical ionization; PID = photoionization detector; SFE = supercritical fluid extraction 
HEXACHLOROBENZENE 253 

7. ANALYTICAL METHODS 

Florisil (Mes et al. 1982), or a combination of columns (Burse et al. 1990). Analysis is usually by 

GC/ECD (Bristol et al. 1982; Burse et al. 1990; EPA 1980c; Mes et al. 1982), although GC coupled with 

photoionization detection (PID) may be used as well (Langhorst and Nestrick 1979). Confirmation by 

GC/MS is recommended (Bristol et al. 1982; Mes et al. 1982). Recovery for all methods is acceptable 

(.70–80%) (Bristol et al. 1982; Burse et al. 1990; Langhorst and Nestrick 1979; Mes et al. 1982); 

precision is also acceptable (#20% relative standard deviation [RSD]) (Bristol et al. 1982; Burse et al. 

1990; Langhorst and Nestrick 1979; Mes et al. 1982). Detection limits are in the low-ppb (ng/g) range 

(Bristol et al. 1982; Burse et al. 1990; Langhorst and Nestrick 1979; Mes et al. 1982). 

Adipose tissue is usually solvent extracted (EPA 1980c; Mes et al. 1982), and the hexachlorobenzene is 

separated from the extracted fat by Florisil column fractionation (Mes et al. 1982). Analysis is by 

GC/ECD (EPA 1980c; Mes et al. 1982). Confirmation by GC/MS (Mes et al. 1982) or a second GC 

column is recommended. Recovery is good (82–86%) (Mes et al. 1982); precision is very good (<10% 

RSD) (Mes et al. 1982). Solvent extraction followed by gel permeation chromatography (GPC) clean-up 

and Florisil column fractionation was utilized for a large adipose tissue monitoring study (EPA 1986f). 

Additional clean-up measures may be required if fractions are not clean enough for capillary GC/MS 

analysis (EPA 1986f). Supercritical fluid extraction (SFE) and treatment with alumina for lipid removal 

have been combined; additional purification was carried out by column chromatography (Djordjevic et al. 

1994). Recovery was 115%, precision 10.5% RSD. Detection limits for all methods are in the low-ppb 

(ng/g) range (Alawi and Ababneh 1991; Djordjevic et al. 1994; EPA 1986f; Mes et al. 1982). 

Few methods are available for monitoring other tissues and fluids. Breast milk has been analyzed with a 

combination of fat separation, column clean-up, and capillary GC/ECD (Abraham et al. 1994). Detection 

limits were 0.4 ng/g fat; other performance data were not reported. Methods for urine (Langhorst and 

Nestrick 1979) and semen (Stachel et al. 1989) have been reported. Both provide good recovery 

(80–84%). A method for feces has been reported, and involves boiling with solvent and clean-up on 

alumina followed by capillary GC/ECD analysis (Abraham et al. 1994). Performance data were not 

reported. 

It is well known that ingestion of hexachlorobenzene can produce porphyria (see Section 3.2.2.2). 

Urinary porphyrins from humans with porphyria cuntanea tarda (PCT) can be analyzed using thin layer 

chromatography (TLC). Separation and estimation of porhyrins are carried out on a TLC plate by 

extraction and esterification of porphyrins, 2-dimensional development, and fluorescent scanning (Miura 

and Torinuki 1977). Other analysis methods for porphyrins include spectrophotometry. Analysis by this 
HEXACHLOROBENZENE 254 

7. ANALYTICAL METHODS 

method is carried out by extraction of porphyrins using an anion exchange column, esterification of 

porphyrins, separation by chromatography, and quanitification spectrophotometrically (Grinstein 1977). 

7.2 ENVIRONMENTAL SAMPLES 

Most environmental analyses have been performed using multiresidue methods involving solvent extract 

of the analytes from the sample matrix, clean-up to remove interfering compounds, determination by GC 

with ECD, and confirmation using an ancillary method such as MS. New methods and technologies are 

evolving, and this has resulted in lower detection limits. For example, detection limits are in the low ppb 

to ppt range for water matrices and the low ppm to ppb range for food. Analytical methods for the 

determination of hexachlorobenzene in environmental samples is given in Table 7-2. 

Atmospheric hexachlorobenzene is usually sampled by pulling a volume of air through an adsorbent trap 

(EPA 1988d, 1988h; Hippelein et al. 1993; Langhorst and Nestrick 1979). A filter may be included in the 

sampling system in order to determine the amount of hexachlorobenzene in particulate (Atlas and Giam 

1981; Brorström-Lundén et al. 1994; Hippelein et al. 1993). Filters and polyurethane foam (PUF) 

adsorbent are Soxhlet extracted (EPA 1988c, 1988h; Hippelein et al. 1993); XAD-2 adsorbent is extracted 

in a Soxhlet apparatus (Hippelein et al. 1993) or by solvent desorption (Langhorst and Nestrick 1979). 

Clean-up on adsorbent columns may be utilized (EPA 1988d; Hippelein et al. 1993). A variety of 

analytical methods are used: GC/ECD (Atlas and Giam 1981; EPA 1988c), capillary GC/ECD 

(Brorström-Lundén et al. 1994; EPA 1988h), GC/PID (Langhorst and Nestrick 1979), and capillary 

GC/MS (Hippelein et al. 1993). Confirmation on a second GC column or by GC/MS is recommended 

(Atlas and Giam 1981; EPA 1988h). Reported recovery is good (82–103%) (EPA 1988h; Langhorst and 

Nestrick 1979); precision is also good (<10%) (Hippelein et al. 1993). Detection limits depend upon the 

amount of air sampled, but may be in the ppb to sub-ppt range (EPA 1988h; Hippelein et al. 1993; 

Langhorst and Nestrick 1979). 

Hexachlorobenzene is usually extracted from water with organic solvents for analysis (EPA 1988e, 

1988f; Munch et al. 1990). Hexachlorobenzene may also be extracted and concentrated by adsorption on 

adsorbent cartridges or disks, with subsequent solvent desorption (EPA 1988a). Clean-up of the extracts 

is usually not necessary; however, methods are available for samples that contain interfering compounds 

(Chan et al. 1994; Driscoll et al. 1991; Garrison and Pellizzari 1987). Analysis is usually by capillary 
Table 7-2. Analytical Methods for Determining Hexachlorobenzene in Environmental 

Sample matrix Preparation method Analytical method 

Sample detection limit 

Percent recover

Air Collection on PUF; Soxhlet extraction; 

cleanup on alumina 

(EPA Method TO-10) GC/ECD 

No data No data 

Ambient air .2,200 m3 collected on GFF and XAD-2; 

Soxhlet extraction; cleanup on layered silica gel; alumina partition 

cap. GC/MS 0.18 pg/m3 

(calculated) 

No data 

Ambient air Collection on XAD-2; solvent desorption GC/PID 70 ppb 95 

Ambient air Collection on PUF; Soxhlet extraction; 

concentration 

dual column megabore GC/ECD or GC/ECD and GC/MS 

5 ng/m3 82–103 

Rain, snow Modified collector; solvent extraction; 

solvent exchange; cleanup on silica gel 

cap. GC/ECD 0.4 ng/L No data 

Drinking water Solid-phase extraction (disk or cartridge) (EPA Method 525.1) 

cap. GC/MS 

0.1–0.2 μg/L 98–109 

Drinking water Solvent extraction; solvent exchange (EPA Method 508) 

cap. GC/ECD; confirmation using second column 

0.077 μg/L (estimated) 

68–82 
Table 7-2. Analytical Methods for Determining Hexachlorobenzene in Environmental Samp

Sample matrix Preparation method Analytical method 

Sample detection limit 

Percent recover

Drinking water Solvent extraction (EPA Method 505) 

GC/ECD, confirmation using second column 

0.003 μg/L 91–100 

Drinking water pH adjustment; concentration on XAD-4; 

cleanup on silica gel 

(Master Scheme) cap. GC/MS 

0.1 μg/L (target) 73 

Groundwater Solvent extraction; solvent exchange (National Pesticide 

Survey Method 2) cap. GC/ECD, confirmation using second column 

0.12 μg/L 96 

River water Centrifugation; chromic acid digestion; 

extraction 

cap. GC/ECD No data 97.5 

Municipal and industrial waste 

Solvent extraction; solvent exchange; optional cleanup on Florisil 

(EPA Method 612) GC/ECD 

0.05 μg/L 95 

Municipal and industrial waste 

pH adjustment; solvent extraction; concentration 

(EPA Method 625) GC/MS 

1.9 μg/L 79 

Waste water, soil, sediments, solid wastes 

Solvent extraction (EPA Method 8410) 

cap. GC/FTIR 

20 μg/L Not 

applicab
Table 7-2. Analytical Methods for Determining Hexachlorobenzene in Environmental Samp

Sample matrix Preparation method Analytical method 

Sample detection limit 

Percent recover

Groundwater, soils, solid wastes 

Various extraction; cleanup methods (EPA Method 8270B) 

cap. GC/MS 

660 μg/kg (soil, sediment); 10 μg/L (ground water) 

72.6 (automaSoxhlet extractio

Soil Solvent extraction; liquid-liquid partition; 

cleanup by sulfuric acid treatment 

GC/ECD No data 98 

Soil Soxhlet and sonication extraction; 

acetylation; solvent extraction; fractionation on silica gel 

dual column cap. GC/ECD 

No data 83–106 

Sediments Microwave extraction centrifugation; 

filtration 

cap. GC/ECD No data 91.7 

Fish tissue Grind with sodium sulfate; 

extract with hexane/acetone

 GC/ECD No data No data 

Fish Homogenization; Soxhlet extraction; GPC 

fractionation; silica gel fractionation 

cap. GC/MS 12.5 ng/g 96 

Fish Maceration; Soxhlet extraction; cleanup 

with sulfuric acid/silica gel 

dual cap. GC/ECD 5 ng/g (lipid basis) 95 
Table 7-2. Analytical Methods for Determining Hexachlorobenzene in Environmental Samp

Sample matrix Preparation method Analytical method 

Sample detection limit 

Percent recover

Fish, aquatic biota 

Homogenization with solvent; solvent exchange; cleanup on Florisil 

cap. GC/ECD, confirmation on second column 

0.01 mg/kg ~94 

Aquatic organisms 

Homogenization; Soxhlet extraction; GPC fractionation; SPE fractionation; solvent exchange 

(USGS method) cap. GC/ECD 

No data 50–75 

Butterfat, fish Isolation on Florisil column; solvent 

partition; partition on Florisil 

GC/ECD No data 95–98 (f

99–104 (butterfa

Fatty foods SFE/SFC (on-line cleanup) cap. GC/ECD 4 ppb 85 

Fatty foods Extraction and pretreatment; Florisil 

cleanup 

(DFG Method S9) GC/ECD; confirmation by TLC 

0.01 mg/kg 90 

Milk Solid phase extraction GC/ECD No data 88–94 

Milk Solvent extraction; solvent partition; 

solvent exchange; GPC cleanup; optional alumina cleanup 

GC/ECD, confirmation on second column 

<0.5 ppb 88–91 
Table 7-2. Analytical Methods for Determining Hexachlorobenzene in Environmental Samp

Sample 

Sample	 detection Percent matrix Preparation method	 Analytical method limit recoverVegetable oils, oil seeds Fruits, vegetables Crops and foods Pine needles 

Sandwich-type extraction fractionation 

Chop and blend; blend with solvent; partition with water; dry Solvent extraction; GPC cleanup; optional silica gel cleanup 

Dry and mince; homogenization; Soxhlet 

extraction; sulfuric acid cleanup; fractionation on Florisil 

GC/ECD 1–2 ppb 80–100 

GC/ECD, confirmation 0.002 ppm 93 

by GC/MS (DFG Method S19) No data >70 

dual GC/ECD cap. GC/ECD	 0.1 ng/g (dry 80–100 

weight) 

cap. = capillary; ECD = electron capture detector; FTIR = Fourier transform infrared spectrometry; GC = gas chromatography; GFF = g

gel permeation chromatography; MS = mass spectrometry; PID = photoionization detector; PUF = polyurethane foam; SFC = supercrit

chromatography; SFE = supercritical fluid extraction; SPE = solid phase extraction; TLC = thin-layer chromatography; USGS = U.S. Ge
HEXACHLOROBENZENE 260 

7. ANALYTICAL METHODS 

GC/ECD (Chan et al. 1994; Driscoll et al. 1991; EPA 1988f; Munch et al. 1990). Confirmation using a 

second method is recommended (EPA 1988e, 1988f; Munch et al. 1990). Capillary GC/MS is also 

utilized for analysis (EPA 1988c; Garrison and Pellizzari 1987). Accuracy ranges from acceptable 

(.60–80%) (EPA 1988a; Garrison and Pellizzari 1987) to excellent (>90%) (Driscoll et al. 1991; EPA 

1988c, 1988e; Munch et al. 1990). Precision is rarely reported; 16% RSD was reported for the Master 

Scheme (Garrison and Pellizzari 1987). Detection limits are in the low- to sub-ppb range (EPA 1988c; 

Garrison and Pellizzari 1987; Munch et al. 1990). Detection limits in the ppt range have been achieved 

by methods utilizing solvent extraction with capillary GC/ECD analysis (Chan et al. 1994; EPA 1988e). 

Waste water is solvent extracted with analysis by GC/ECD (EPA 1984e) or GC/MS (EPA 1988e). 

Reported recovery is good (79–95%) (EPA 1984e, 1984f). Detection limits are in the low-ppb range, 

with lower detection limits reported for the GC/ECD analysis (EPA 1984e). 

Soxhlet or sonication extraction is most commonly used to extract hexachlorobenzene from solid matrices 

such as soils and sediments, and wastes (EPA 1984e; Ojala 1993). Solvent extraction (Waliszewski and 

Szymczynski 1985) and microwave extraction techniques (Onuska and Terry 1993) may be used as well. 

Clean-up is usually required for the extracts (EPA 1994a; Ojala 1993; Waliszewski and Szymczynski 

1985), with subsequent analysis by GC/ECD (Waliszewski and Szymczynski 1985), capillary GC/ECD 

(Ojala 1993; Onuska and Terry 1993), or capillary GC/MS (EPA 1994a). Reported recovery is good 

(73–106%) (EPA 1994a; Ojala 1993; Onuska and Terry 1993; Waliszewski and Szymczynski 1985). 

Precision, where reported, is acceptable (#20% RSD) (EPA 1984e, 1984f; Ojala 1993). Little 

information is available on detection limits. Detection limits of 660 μg/kg (ppb) have been reported for 

automated Soxhlet extraction with capillary GC/MS analysis (EPA 1994a). 

Fish and aquatic organisms are homogenized, then extracted with solvent (Miskiewicz and Gibbs 1994; 

Oliver and Nichol 1982b), isolated on Florisil columns (Bong 1975), or Soxhlet extracted (Rahman et al. 

1993; Shan et al. 1994; Tiernan et al. 1990). Clean-up is usually necessary to remove lipids and 

interfering substances (Bong 1975; Miskiewicz and Gibbs 1994; Rahman et al. 1993; Shan et al. 1994; 

Tiernan et al. 1990). Capillary GC/ECD analysis is used most often (Miskiewicz and Gibbs 1994; 

Rahman et al. 1993; Shan et al. 1994). Capillary GC/MS (Tiernan et al. 1990) and GC/ECD (Bong 1975; 

Oliver and Nichol 1982b) are also utilized. Reported recovery ranges from moderate (50–75%) (Shan et 

al. 1994) to excellent (>90%) (Bong 1975; Miskiewicz and Gibbs 1994; Rahman et al. 1993; Tiernan et 

al. 1990). Precision is usually not reported; however, 4–6% RSD has been achieved (Shan et al. 1994). 

Detection limits, where reported, are in the low-ppb range (ng/g) (Miskiewicz and Gibbs 1994; Rahman 

et al. 1993; Tiernan et al. 1990). 
HEXACHLOROBENZENE 261 

7. ANALYTICAL METHODS 

Fatty foods, including milk, have been extracted with solvent to remove the fat, then cleaned up to 

separate the hexachlorobenzene from the fat (AOAC 1990; Bong 1975; Thier and Zeumer 1987b; Trotter 

and Dickerson 1993). Newer methods for combined separation and clean-up are supercritical fluid 

techniques (Nam and King 1994), solid-phase extraction (SPE) (Manes et al. 1993), and a sandwich 

system (Seidel and Linder 1994). Analysis is by GC/ECD (AOAC 1990; Bong 1975; Manes et al. 1993; 

Seidel and Linder 1993; Thier and Zeumer 1987b; Trotter and Dickerson 1993). Confirmation on a 

second GC column (Trotter and Dickerson 1993) or by thin-layer chromatography (TLC) (Thier and 

Zeumer 1987b) is recommended. Capillary GC/ECD has also been utilized (Nam and King 1994). 

Reported recoveries are good (>80%) (Bong 1975; Manes et al. 1993; Nam and King 1994; Seidel and 

Linder 1993; Trotter and Dickerson 1993). Precision, where reported, is very good (<15% RSD) (Bong 

1975; Nam and King 1994; Thier and Zeumer 1987b; Trotter and Dickerson 1993). Limit of detection, 

where reported, is in the low-ppb (ng/g) range (Nam and King 1994; Seidel and Linder 1993; Thier and 

Zeumer 1987b; Trotter and Dickerson 1993). 

Fruits, vegetables, and crops are blended, solvent extracted, and then cleaned up and fractionated 

(Pylypiw 1993; Thier and Zeumer 1987a). Capillary GC/ECD is the analytical method. Recovery is 

acceptable (>70%) (Pylypiw 1993; Thier and Zeumer 1987a). Precision was not reported. The reported 

detection limit is 2 ppb (Pylypiw 1993). 

7.3 ADEQUACY OF THE DATABASE 

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with the 

Administrator of EPA and agencies and programs of the Public Health Service) to assess whether 

adequate information on the health effects of hexachlorobenzene is available. Where adequate 

information is not available, ATSDR, in conjunction with the National Toxicology Program (NTP), is 

required to assure the initiation of a program of research designed to determine the health effects (and 

techniques for developing methods to determine such health effects) of hexachlorobenzene. 

The following categories of possible data needs have been identified by a joint team of scientists from 

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would 

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean 

that all data needs discussed in this section must be filled. In the future, the identified data needs will be 

evaluated and prioritized, and a substance-specific research agenda will be proposed. 
HEXACHLOROBENZENE 262 

7. ANALYTICAL METHODS 

7.3.1 Identification of Data Needs 

Methods for Determining Biomarkers of Exposure and Effect. Methods exist for measuring 

hexachlorobenzene in blood (Bristol et al. 1982; Burse et al. 1990; Langhorst and Nestrick 1979; Mes et 

al. 1982) and adipose tissue (Alawi and Ababneh 1991; Djordjevic et al. 1994; EPA 1980c, 1986f; Mes et 

al. 1982). The methods for blood and adipose are sensitive (low-ppb range), but improved accuracy is 

needed for blood analysis. The data on determination of hexachlorobenzene in urine, breast milk, and 

tissues are limited, and the methods may not be sufficiently sensitive. Methods that could be used to 

measure low levels in human tissues would be useful for determining the relationship between chronic 

low-level exposure and the effects observed in specific tissues. Improved methods to detect phenolic 

metabolites are not needed since these metabolites are not unique to hexachlorobenzene. Representative 

methods for determining pentachlorophenol and other phenolic metabolites using GC/ECD and GC/MS 

are shown in Table 7-3. 

Biomarkers for effects of hexachlorobenzene are porphyric symptoms and increased gamma-glutamic 

transferase activity. Since these effects are also indicative of exposure to other toxicants, additional 

studies are needed for more specific biomarkers for effects of hexachlorobenzene exposure. 

Methods for Determining Parent Compounds and Degradation Products in Environmental 

Media. Methods for determining hexachlorobenzene in air (EPA 1988d, 1988h; Hippelein et al. 1993; 

Langhorst and Nestrick 1979) and water (Chan et al. 1994; EPA 1988c, 1988e, 1988f; Garrison and 

Pellizzari 1987), the media of most concern for human exposure, are reliable, but may not be sensitive 

enough to measure background levels in the environment. Limited performance data are available for 

methods for soil and other solid media. In addition, there is insufficient performance information for 

methods for determining hexachlorobenzene in media such as shellfish, fish, and plants. Some exposure 

to hexachlorobenzene may occur via ingestion of food and standardized methods for foods are needed. 

Methods with sufficient sensitivity for measuring background levels in foods would be helpful as well. 
HEXACHLOROBENZENE 263 

7. ANALYTICAL METHODS 

Table 7-3. Analytical Methods for Determining Biomarkers of
 

Hexachlorobenzene
 

Sample matrix 

Preparation 

method 

Analytical 

method 

Sample 

detection 

limit 

Percent 

recovery Reference 

Blood 

(pentachlorophenol) 

pH adjustment, 

solvent 

extraction, 

derivatization 

GC/ECD 10 ppb 92 EPA 1980c 

Urine 

(chlorinated 

phenol 

metabolites) 

Hydrolysis; 

solvent 

extraction, 

derivatization 

GC/ECD, 

confirmation 

by GC/MS 

No data >90 (PCP); 

most other 

metabolites >80 

EPA 1980c 

ECD = electron capture detector; GC = gas chromatography; MS = mass spectrometry; PCP = pentachlorophenol 
HEXACHLOROBENZENE 264 

7. ANALYTICAL METHODS 

7.3.2 Ongoing Studies 

The Environmental Health Laboratory Sciences Division of the National Center for Environmental Health 

and Injury Control, Centers for Disease Control and Prevention, is developing methods for the analysis of 

pentachlorophenol and other phenolic compounds in urine. These methods use high resolution GC and 

magnetic sector MS, which gives detection limits in the low ppt range. 

HEXACHLOROBENZENE 266 

8. REGULATIONS AND ADVISORIES
 

The international, national, and state regulations and guidelines regarding hexachlorobenzene in air, 

water, and other media are summarized in Table 8-1. 

ATSDR has derived an acute oral MRL of 8x10-3 mg/kg/day on the basis of developmental neurotoxic 

effects in rats. This MRL is based on a LOAEL of 2.5 mg/kg/day for hyperactivity in rat offspring whose 

mothers were exposed to hexachlorobenzene for 4 days prior to mating (Goldey and Taylor 1992). 

An intermediate oral MRL of 1x10-4 mg/kg/day, based on reproductive effects in monkeys (Babineau 

et al. 1991; Bourque et al. 1995; Jarrell et al. 1993) has also been derived for hexachlorobenzene by 

ATSDR. This MRL is based on degenerative changes in the ovaries of female monkeys exposed to 

hexachlorobenzene doses of $0.01 mg/kg/day for 90 days. 

ATSDR has also derived a chronic oral MRL of 5x10-5 mg/kg/day, based on a LOAEL of 

0.016 mg/kg/day for peribiliary lymphocytosis and fibrosis of the liver in adult F1 generation male rats 

fed hexachlorobenzene for 130 weeks in a 2-generation study by Arnold et al. (1985). 

The EPA oral reference dose for hexachlorobenzene (IRIS 2002) is 8x10-4 mg/kg/day based on liver 

effects in the Arnold et al. (1985) rat study. No reference concentration exists for the compound. These 

EPA assessments are currently undergoing re-evaluation. 

EPA has classified hexachlorobenzene in weight-of-evidence Group B2 as a probable human carcinogen 

(IRIS 2001). EPA derived an oral slope factor of 1.6 per (mg/kg)/day and an inhalation unit risk of 

4.6x10-4 per (μg/m3

) based on hepatocellular carcinoma in rats exposed orally. This EPA assessment is 

currently being re-evaluated. IARC classifies the chemical as 2B, possibly carcinogenic to humans 

(IARC 2001). The National Toxicology Program (2001) concluded that hexachlorobenzene is reasonably 

anticipated to be a human carcinogen. 

Hexachlorobenzene is on the list of chemicals appearing in "The Emergency Planning and Community 

Right-to-Know Act of 1986" (EPCRA) (EPA 2001a). Section 313 of Title III of EPCRA requires owners 

and operators of certain facilities that manufacture, import, process, or otherwise use the chemicals on this 

list to report annually their release of those chemicals to any environmental media. 
HEXACHLOROBENZENE 267 

8. REGULATIONS AND ADVISORIES 

Although no Occupational Safety and Health Administration standards exist for hexachlorobenzene, the 

American Conference of Governmental Industrial Hygienists has set a threshold limit value (8-hour time 

weighted average) of 0.002 mg/m3

 (ACGIH 2001), based on a route-to-route extrapolation from an oral 

study in Rhesus monkeys (Rozman et al. 1978). 

Hexachlorobenzene is regulated by the Clean Water Effluent Guidelines as stated in Title 40, 

Section 400–475, of the Code of Federal Regulations. For each point source category, hexachlorobenzene may be regulated as one of a group of chemicals that are controlled as Total Toxic Organics, or 

may have a specific Regulatory Limitation, or may have a Zero Discharge Limitation. The point source 

categories for which hexachlorobenzene is controlled as a Total Toxic Organic are electroplating (EPA 

2001b), metal finishing (EPA 2001c), and coil coating. The point source categories for which 

hexachlorobenzene has a specific Regulatory Limitation include primary rare earth metals (EPA 2001d) 

and organic chemicals, plastics, and synthetic fibers (EPA 2001e). 

The Resource Conservation and Recovery Act (RCRA) identifies hexachlorobenzene as a hazardous 

waste in two ways: (1) when it exceeds a toxicity characteristic leaching procedure test concentration of 

0.13 mg/L (EPA 2000p), and (2) when discarded as a commercial product, off-spec species, container 

residue, or spill residue (EPA 2001i). 
HEXACHLOROBENZENE 268 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to Hexachlorobenzene 

Agency Description	 Information Reference 

INTERNATIONAL 

Guidelines: 

IARC	 Carcinogenicity classification Group 2Ba IARC 2001 

WHO	 Drinking water guideline 1 μg/L WHO 1996 

Total daily intake in humans WHO 1997 

Non-cancer effects 0.17 μg/kg b.w./day 

Neoplastic effects 0.16 μg/kg b.w./day 

NATIONAL 

Regulations and 

Guidelines: 

a. Air 

ACGIH	 TLV (8-hour TWA) 0.002 mg/m3 ACGIH 2001 

NIOSH	 REL No data 

OSHA	 PEL No data 

USC	 Listed as a hazardous air USC 2001
 

pollutant 42 USC 7412
 

b. Water 

EPA Drinking water standard 0.001 ppm	 EPA 2001m 

40CFR141.32(e)(68) 

MCL 0.001 mg/L	 EPA 2001k 

40CFR141.61(c) 

MCLG Zero	 EPA 2001l 

40CFR141.50(a)(22) 

Effluent guidelines and Total toxic organics EPA 2001b 

standards—electroplating point 40CFR413.02 

source category 

Effluent guidelines and Total toxic organics EPA 2001c 

standards—metal finishing point 40CFR433.11 

source category 

Effluent guidelines and New source EPA 2001d 

standards—nonferrous metals performance 40CFR421.274 

manufacturing point source standards for the 

category primary rare earth 

metals subcategory 

Effluent guidelines and EPA 2001e 

standards—organic chemicals, 40CFR414 

plastics, and synthetic fibers 
HEXACHLOROBENZENE 269 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to Hexachlorobenzene 

(continued) 

Agency Description Information Reference 

NATIONAL (cont.) 

EPA Groundwater monitoring 

Suggested methods 

8120 

8270 

PQL 

0.5 μg/L 

10 μg/L 

EPA 2001f 

40CFR264, 

Appendix IX 

Land disposal restrictions; 

universal treatment standards 

Waste water 

Non-waste water 

0.055 mg/L2 

10 mg/kg3 

EPA 2001j 

40CFR268.48(a) 

c. Food 

USDA Labeling of treated seed with 

commonly accepted chemical 

(generic) name of substance 

USDA 2001 

7CFR201.31a(b) 

d. Other 

ACGIH Carcinogenicity classification A3b ACGIH 2001 

EPA Carcinogenicity classification 

Oral slope factor 

Inhalation unit risk 

RfD 

RfC 

B2c 

1.6 per (mg/kg)/day 

4.6x10-4 per (μg/m3

) 

8x10-4 mg/kg/day 

No data 

IRIS 2001 

Health-based limits for exclusion 

of waste-derived residues— 

concentration limits for residues 

2x10-4 mg/kg EPA 2001h 

40CFR266, 

Appendix VII 

Identification and listing of 

hexachlorobenzene as a 

hazardous waste—hazardous 

waste number 

U127 EPA 2001i 

40CFR261.33(f) 

Maximum concentration of 

contaminants for the toxicity 

characteristic—regulatory level 

0.13 mg/L EPA 2001p 

40CFR261.24 
HEXACHLOROBENZENE 270 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to Hexachlorobenzene 

(continued) 

Agency Description Information Reference 

NATIONAL (cont.) 

EPA Reportable quantity regarded as 

a CERCLA hazardous 

substance under Section 307(a) 

of the Clean Water Act, 

Section 112 of the Clean Air Act, 

and Section 3001 of RCRA 

10 pounds EPA 2001a 

40CFR302.4 

Hazardous waste management 

Unit risk 

Risk specific dose 

4.9x10-4 μg/m3 

2.0x10-2 μg/m3 

EPA 2001g 

40CFR266, 

Appendix V 

Toxic chemical release 

reporting; community right-toknow—alternate reporting 

threshold for PBT compounds 

10 pounds EPA 2001o 

40CFR372.27 

Toxic chemical release 

reporting; community right-toknow—effective date of 

reporting 

01/01/87 EPA 2001n 

40CFR372.65 

NTP Carcinogenicity classification Reasonably 

anticipated to be a 

human carcinogen 

NTP 2001 

STATE 

a. Air 

Idaho Acceptable ambient 

concentration for a carcinogen 

Emissions level 

2x10-3 μg/m3 

1.3x10-5 pounds/hour 

ID Dept. of Health 

and Welfare 1999 

Washington Acceptable source impact levels 

(at 10-6 risk), annual average 

0.0022 μg/m3 WA Dept. of Ecology 

1998 

Wisconsin Hazardous air contaminants 

without acceptable ambient 

concentrations requiring 

application of best available 

control technology 

25 pounds/year2 WI Dept. of Natural 

Resources 1997 

b. Water 

Alaska MCL 0.001 mg/L AK Dept. of Environ. 

Conservation 1999 

Arizona Drinking water guideline 0.02 ug/L HSDB 2001 
HEXACHLOROBENZENE 271 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to Hexachlorobenzene 

(continued) 

Agency Description Information Reference 

STATE (cont.) 

California MCL 0.001 mg/L CA Dept. of Health 

Services 2000 

Colorado Groundwater standard 1 μg/L CO Dept. of Public 

Health and Environ. 

1999 

Hawaii MCL applying to community and 

non-transient, non-community 

water systems 

0.001 mg/L HI Dept. of Health 

1999a 

Hawaii Toxic pollutant standardse 

Freshwater 

Acute 

Chronic 

Saltwater 

Acute 

Chronic 

Fish consumption 

No standard 

No standard 

No standard 

No standard 

2.4x10-4 μg/L 

HI Dept. of Health 

1999b 

Kansas Water quality standards 

Aquatic life 

Acute 

Chronic 

Public health 

Food procurement 

Domestic water supply 

6.0 mg/L 

3.7 mg/L 

7.4x10-4 mg/L 

1 mg/L 

KS Dept. of Health 

and Environ. 1999 

Maine Drinking water guideline 0.2 ug/L HSDB 2001 

Minnesota Drinking water guideline 0.2 ug/L HSDB 2001 

New Jersey Groundwater quality criteria 

PQL 

0.02 μg/L 

10 μg/L 

NJ Dept. of Environ. 

Protection 1993 

South Dakota MCL for drinking water 0.001 mg/L SD Dept. of Environ. 

Natural Resources 

1998 

c. Food No data 

d. Other 

California Proposition 65 chemical; known 

to the state to cause cancer and 

developmental toxicity 

Any manufacturer, 

packager, or 

producer required to 

add specific warnings 

to products and 

shipments 

BLR 2002 
HEXACHLOROBENZENE 272 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to Hexachlorobenzene 

(continued) 

Agency Description Information Reference 

STATE (cont.) 

Florida Toxic substance A toxic substance is BLR 2002 

present in any 

mixture if it is 1% or 

more of the mixture 

Massachusetts Hazardous substance Carcinogen, 

teratogen, and extraordinary hazardous 

BLR 2002 

Minnesota Hazardous substance Potential hazard from BLR 2002 

absorption through 

skin contact 

New Jersey Right to know hazardous 

substance; required to submit 

surveys listing the hazardous 

substances present at their 

facilities in quantities 

>500 pounds; and report their 

inventories of any chemical that 

requires a MSDS and is present 

on site in quantities 

>10,000 pounds 

BLR 2002 

New York Hazardous substance— BLR 2002 

reportable quantities 

Air 

Land 

10 pounds 

1 pound 

a

Group 2B: possibly carcinogenic to humans

b

A3: confirmed animal carcinogen with unknown relevance to humans 

c

B2: probable human carcinogen

d

Indicates value derived using an oral cancer potency factor as a surrogate. 

ACGIH = American Conference of Governmental Industrial Hygienists; ATSDR = Agency for Toxic Substances 

and Disease Registry; BLR = Business & Legal Reports, Inc.; b.w. = body weight; CERCLA = Comprehensive 

Environmental Response Compensation and Liability Act; CFR = Code of Federal Regulations; 

EPA = Environmental Protection Agency; HSDB = Hazardous Substances Data Bank; IARC = International 

Agency for Research on Cancer; IRIS = Integrated Risk Information System; MCL = maximum contaminant level; 

MCLG = maximum contaminant level goal; MRLs = Minimal Risk Levels; MSDS = Material Safety Data Sheets; 

NIOSH = National Institute for Occupational Safety and Health; NTP = National Toxicology Program; 

OSHA = Occupational Safety and Health Administration; PEL = permissible exposure limit; PQL = practical 

quantitation limit; RCRA = Resource Conservation and Recovery Act; REL = recommended exposure limit; 

RfC = inhalation reference concentration; RfD = oral reference dose; TLV = threshold limit value; TWA = timeweighted average; USC = United States Code; USDA = United States Department of Agriculture; WHO = World 

Health Organization 

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EPA. 1982c. Steam electric power generating point source category. U.S. Environmental Protection 

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EPA. 1982d. Effects of selected organic drinking water contaminants on male reproduction. Research 

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EPA. 1982e. Treatability manual: Volume I. Treatability Data. Washington, DC: U.S. Environmental 

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EPA. 1982f. Preliminary quantitative usage analysis of picloram. Washington, DC: U.S. Environmental 

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*EPA. 1983a. EPA administered permit programs: The national pollutant discharge elimination system. 

Code of Federal Regulations. 40 CFR 122. 

EPA. 1983b. Metal finishing point source category. U.S. Environmental Protection Agency. Code of 

Federal Regulations. 40 CFR 433.11. 

EPA. 1983c. Reportable quantity document for benzene, hexachloro-. External review draft. 

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EPA. 1983d. List of chemicals produced by affected facilities. U.S. Environmental Protection Agency. 

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*EPA. 1984a. Primary rare earth metals. U.S. Environmental Protection Agency. Code of Federal 

Regulations. 40 CFR 421. 
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*EPA. 1984b. Health effects assessment for hexachlorobenzene. Cincinnati, OH: U.S. Environmental 

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EPA. 1984c. Nonindustrial sources of potentially toxic substances and their applicability to source 

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EPA. 1984d. Guidelines establishing test procedures for the analysis of pollutants. U.S. Environmental 

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*EPA. 1984e. Methods for organic chemical analysis of municipal and industrial wastewater. 

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*EPA. 1984f. Methods for organic chemical analysis of municipal and industrial wastewater. 

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EPA. 1985a. Health assessment document for chlorinated benzenes: Final Report. Cincinnati, OH: U.S. 

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*EPA. 1985c. Standards for the management of specific hazardous wastes and specific types of 

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*EPA. 1985g. Environmental exposure to hexachlorobenzene in the United States. Washington, DC: 

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EPA. 1985. Preliminary quantitative usage analysis of PCNB (056502). Washington, DC: U.S. 

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*EPA. 1986a. Method 8250: Gas chromatography/mass spectrometry for semivolatile organics: Packed 

column technique. In: Test methods for evaluating solid waste. 3rd ed. SW-846. Washington, DC: 

U.S. Environmental Protection Agency, Office of Solid Waste and Emergency Response. 

EPA. 1986b. Method 8270: Gas chromatography/mass spectrometry for semivolatile organics: 

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Washington, DC: U.S. Environmental Protection Agency, Office of Solid Waste and Emergency 

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*EPA. 1986c. Method 8410: Capillary column analysis of semivolatile organic compounds by gas 

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EPA. 1986d. Land disposal restrictions. U.S. Environmental Protection Agency. Code of Federal 

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*EPA. 1986e. Exposure assessment for hexachlorobenzene. Washington, DC: U.S. Environmental 

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*EPA. 1986f. Broad scan analysis of the FY82 national human adipose tissue survey specimens: 

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*EPA. 1987a. Organic chemicals, plastics, and synthetic fibers. U.S. Environmental Protection Agency. 

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EPA. 1987f. Reference dose (RfD): Description and use in health risk assessments. Vol. I, Appendix A: 

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*EPA. 1987g. Measurement of hydrolysis rate constants for evaluation of hazardous waste land 

disposal. Vol 2. Washington, DC: U.S. Environmental Protection Agency. EPA-600/53-87-/019. 

*EPA. 1988a. Drinking water criteria document for hexachlorobenzene. Cincinnati, OH: U.S. 

Environmental Protection Agency, Environmental Criteria and Assessment Office, Office of Health and 

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EPA. 1988b. Toxic chemical release reporting: Community right-to-know. U.S. Environmental 

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*EPA. 1988c. Methods for the determination of organic compounds in drinking water. Cincinnati, OH: 

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*EPA. 1988d. Compendium of methods for the determination of toxic organic compounds in ambient 

air. Research Triangle Park, NC: U.S. Environmental Protection Agency, Quality Assurance Division, 

Environmental Monitoring Systems Laboratory. EPA-600/4-89-017. 

*EPA. 1988e. Method 505: Method for the determination of organic compounds in drinking water. 

Washington, DC: U.S. Environmental Protection Agency. EPA 600/4-88/039. 

*EPA. 1988f. Method 508: Method for the determination of organic compounds in drinking water. 

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EPA. 1988g. The emergency planning and community right-to-know act of 1986. Washington, DC: 

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*EPA. 1988h. Analytical and sampling methods of the Nonoccupational Pesticide Exposure Study 

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*EPA. 1989a. U.S. Environmental Protection Agency: Part II. Federal Register. 54:1056-1120. 

EPA. 1989b. U.S. Environmental Protection Agency. Federal Register 54 FR 33418. 

EPA. 1989c. Mortality among Dow Chemical employees of a Texas operations per-tete manufacturing 

plant. Washington, DC: U.S. Environmental Protection Agency. OTS0516752. 

*EPA. 1990. Interim methods for development of inhalation reference concentrations. Washington DC: 

U.S. Environmental Protection Agency. EPA/600/8-90/066A. 

EPA. 1991. Criteria for municipal solid waste landfills (Eff. 10-9-93). Washington, DC: U.S. 

Environmental Protection Agency. Code of Federal Regulations. 40 CFR 258. 

EPA. 1992a. Drinking water regulations and health advisories. Washington, DC: U.S. Environmental 

Protection Agency, Office of Water. 

*EPA. 1992b. “National study of chemical residues in fish.” and “Guidance for assessing chemical 

contaminants data for use in fish advisories.” 2 Volumes. Washington, DC: U.S. Environmental 

Protection Agency, Office of Water, Office of Science and Technology. EPA-823-R-92-008a and -008b. 

EPA. 1992c. General pretreatment regulations for existing and new sources of pollution. U.S. 

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EPA. 1992d. Initial submission: Letter submitting studies on dibromohexachlorooctahydrodimethylnapthalene, decachlorooctahydrodimethanonaphthalene, and Hexabromodiphenylamine (final 

report). Washington, DC: U.S. Environmental Protection Agency. OTS0537064. 

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Hexachlorobenzene; technical DAC 839, dacthal; pure dacthal; DAC 1209; DAC 976 W-letter 022192. 

Washington, DC: U.S. Environmental Protection Agency. OTS0537065. 
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*EPA. 1993a. Guidance for assessing chemical contaminant data for use in fish advisories. Vol. 1. Fish 

sampling and analysis. Washington DC: U.S. Environmental Protection Agency, Office of Water. 

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*EPA. 1994a. Test methods for evaluating solid waste. Volume IB: Laboratory Manual 

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*EPA. 1994b. Hexachlorobenzene carcinogenicity summary table. Office of health and Environmental 

Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH. Washington DC: Office of 

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*EPA. 1994c. Personal communication to Dr. Patricia Cunningham, Research Triangle Institute, 

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*EPA. 1995b. Guidance for assessing chemical contaminant data for use in fish advisories. Vol 1: Fish 

sampling and analysis. Second Edition. Washington, DC: U.S. Environmental Protection Agency, 

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*EPA. 1995c. Toxic chemical release inventory. Reporting form R and instructions. Washington DC: 

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*EPA. 1997a. Automated Form R for Windows: User’s guide (RY97). Washington, DC: U.S. 

Environmental Protection Agency, Office of Pollution Prevention and Toxics. 

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*EPA. 1999a. Chemicals and chemical categories to which this part applies. Washington, DC: U.S. 

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HEXACHLOROBENZENE 346 

10. GLOSSARY 

Absorption—The taking up of liquids by solids, or of gases by solids or liquids. 

Acute Exposure—Exposure to a chemical for a duration of 14 days or less, as specified in the 

Toxicological Profiles. 

Adsorption—The adhesion in an extremely thin layer of molecules (as of gases, solutes, or liquids) to the 

surfaces of solid bodies or liquids with which they are in contact. 

Adsorption Coefficient (Koc)—The ratio of the amount of a chemical adsorbed per unit weight of 

organic carbon in the soil or sediment to the concentration of the chemical in solution at equilibrium. 

Adsorption Ratio (Kd)—The amount of a chemical adsorbed by a sediment or soil (i.e., the solid phase) 

divided by the amount of chemical in the solution phase, which is in equilibrium with the solid phase, at a 

fixed solid/solution ratio. It is generally expressed in micrograms of chemical sorbed per gram of soil or 

sediment. 

Benchmark Dose (BMD)—Usually defined as the lower confidence limit on the dose that produces a 

specified magnitude of changes in a specified adverse response. For example, a BMD10 would be the 

dose at the 95% lower confidence limit on a 10% response, and the benchmark response (BMR) would be 

10%. The BMD is determined by modeling the dose response curve in the region of the dose response 

relationship where biologically observable data are feasible. 

Benchmark Dose Model—A statistical dose-response model applied to either experimental toxicological 

or epidemiological data to calculate a BMD. 

Bioconcentration Factor (BCF)—The quotient of the concentration of a chemical in aquatic organisms 

at a specific time or during a discrete time period of exposure divided by the concentration in the 

surrounding water at the same time or during the same period. 

Biomarkers—Broadly defined as indicators signaling events in biologic systems or samples. They have 

been classified as markers of exposure, markers of effect, and markers of susceptibility. 

Cancer Effect Level (CEL)—The lowest dose of chemical in a study, or group of studies, that produces 

significant increases in the incidence of cancer (or tumors) between the exposed population and its 

appropriate control. 

Carcinogen—A chemical capable of inducing cancer. 

Case-Control Study—A type of epidemiological study which examines the relationship between a 

particular outcome (disease or condition) and a variety of potential causative agents (such as toxic 

chemicals). In a case-controlled study, a group of people with a specified and well-defined outcome is 

identified and compared to a similar group of people without outcome. 

Case Report—Describes a single individual with a particular disease or exposure. These may suggest 

some potential topics for scientific research but are not actual research studies. 
HEXACHLOROBENZENE 347 

10. GLOSSARY 

Case Series—Describes the experience of a small number of individuals with the same disease or 

exposure. These may suggest potential topics for scientific research but are not actual research studies. 

Ceiling Value—A concentration of a substance that should not be exceeded, even instantaneously. 

Chronic Exposure—Exposure to a chemical for 365 days or more, as specified in the Toxicological 

Profiles. 

Cohort Study—A type of epidemiological study of a specific group or groups of people who have had a 

common insult (e.g., exposure to an agent suspected of causing disease or a common disease) and are 

followed forward from exposure to outcome. At least one exposed group is compared to one unexposed 

group. 

Cross-sectional Study—A type of epidemiological study of a group or groups which examines the 

relationship between exposure and outcome to a chemical or to chemicals at one point in time. 

Data Needs—Substance-specific informational needs that if met would reduce the uncertainties of human 

health assessment. 

Developmental Toxicity—The occurrence of adverse effects on the developing organism that may result 

from exposure to a chemical prior to conception (either parent), during prenatal development, or 

postnatally to the time of sexual maturation. Adverse developmental effects may be detected at any point 

in the life span of the organism. 

Dose-Response Relationship—The quantitative relationship between the amount of exposure to a 

toxicant and the incidence of the adverse effects. 

Embryotoxicity and Fetotoxicity—Any toxic effect on the conceptus as a result of prenatal exposure to 

a chemical; the distinguishing feature between the two terms is the stage of development during which the 

insult occurs. The terms, as used here, include malformations and variations, altered growth, and in utero 

death. 

Environmental Protection Agency (EPA) Health Advisory—An estimate of acceptable drinking water 

levels for a chemical substance based on health effects information. A health advisory is not a legally 

enforceable federal standard, but serves as technical guidance to assist federal, state, and local officials. 

Epidemiology—Refers to the investigation of factors that determine the frequency and distribution of 

disease or other health-related conditions within a defined human population during a specified period. 

Genotoxicity—A specific adverse effect on the genome of living cells that, upon the duplication of 

affected cells, can be expressed as a mutagenic, clastogenic or carcinogenic event because of specific 

alteration of the molecular structure of the genome. 

Half-life—A measure of rate for the time required to eliminate one half of a quantity of a chemical from 

the body or environmental media. 

Immediately Dangerous to Life or Health (IDLH)—The maximum environmental concentration of a 

contaminant from which one could escape within 30 minutes without any escape-impairing symptoms or 

irreversible health effects. 
HEXACHLOROBENZENE 348 

10. GLOSSARY 

Incidence—The ratio of individuals in a population who develop a specified condition to the total 

number of individuals in that population who could have developed that condition in a specified time 

period. 

Intermediate Exposure—Exposure to a chemical for a duration of 15–364 days, as specified in the 

Toxicological Profiles. 

Immunologic Toxicity—The occurrence of adverse effects on the immune system that may result from 

exposure to environmental agents such as chemicals. 

Immunological Effects—Functional changes in the immune response. 

In Vitro—Isolated from the living organism and artificially maintained, as in a test tube. 

In Vivo—Occurring within the living organism. 

Lethal Concentration(LO) (LCLO)—The lowest concentration of a chemical in air which has been 

reported to have caused death in humans or animals. 

Lethal Concentration(50) (LC50)—A calculated concentration of a chemical in air to which exposure for a 

specific length of time is expected to cause death in 50% of a defined experimental animal population. 

Lethal Dose(LO) (LDLO)—The lowest dose of a chemical introduced by a route other than inhalation that 

has been reported to have caused death in humans or animals. 

Lethal Dose(50) (LD50)—The dose of a chemical which has been calculated to cause death in 50% of a 

defined experimental animal population. 

Lethal Time(50) (LT50)—A calculated period of time within which a specific concentration of a chemical 

is expected to cause death in 50% of a defined experimental animal population. 

Lowest-Observed-Adverse-Effect Level (LOAEL)—The lowest exposure level of chemical in a study, 

or group of studies, that produces statistically or biologically significant increases in frequency or severity 

of adverse effects between the exposed population and its appropriate control. 

Lymphoreticular Effects—Represent morphological effects involving lymphatic tissues such as the 

lymph nodes, spleen, and thymus. 

Malformations—Permanent structural changes that may adversely affect survival, development, or 

function. 

Minimal Risk Level (MRL)—An estimate of daily human exposure to a hazardous substance that is 

likely to be without an appreciable risk of adverse noncancer health effects over a specified route and 

duration of exposure. 

Modifying Factor (MF)—A value (greater than zero) that is applied to the derivation of a minimal risk 

level (MRL) to reflect additional concerns about the database that are not covered by the uncertainty 

factors. The default value for a MF is 1. 
HEXACHLOROBENZENE 349 

10. GLOSSARY 

Morbidity—State of being diseased; morbidity rate is the incidence or prevalence of disease in a specific 

population. 

Mortality—Death; mortality rate is a measure of the number of deaths in a population during a specified 

interval of time. 

Mutagen—A substance that causes mutations. A mutation is a change in the DNA sequence of a cell’s 

DNA. Mutations can lead to birth defects, miscarriages, or cancer. 

Necropsy—The gross examination of the organs and tissues of a dead body to determine the cause of 

death or pathological conditions. 

Neurotoxicity—The occurrence of adverse effects on the nervous system following exposure to a 

chemical. 

No-Observed-Adverse-Effect Level (NOAEL)—The dose of a chemical at which there were no 

statistically or biologically significant increases in frequency or severity of adverse effects seen between 

the exposed population and its appropriate control. Effects may be produced at this dose, but they are not 

considered to be adverse. 

Octanol-Water Partition Coefficient (Kow)—The equilibrium ratio of the concentrations of a chemical 

in n-octanol and water, in dilute solution. 

Odds Ratio (OR)—A means of measuring the association between an exposure (such as toxic substances 

and a disease or condition) which represents the best estimate of relative risk (risk as a ratio of the 

incidence among subjects exposed to a particular risk factor divided by the incidence among subjects who 

were not exposed to the risk factor). An odds ratio of greater than 1 is considered to indicate greater risk 

of disease in the exposed group compared to the unexposed. 

Organophosphate or Organophosphorus Compound—A phosphorus containing organic compound 

and especially a pesticide that acts by inhibiting cholinesterase. 

Permissible Exposure Limit (PEL)—An Occupational Safety and Health Administration (OSHA) 

allowable exposure level in workplace air averaged over an 8-hour shift of a 40-hour workweek. 

Pesticide—General classification of chemicals specifically developed and produced for use in the control 

of agricultural and public health pests. 

Pharmacokinetics—The science of quantitatively predicting the fate (disposition) of an exogenous 

substance in an organism. Utilizing computational techniques, it provides the means of studying the 

absorption, distribution, metabolism and excretion of chemicals by the body. 

Pharmacokinetic Model—A set of equations that can be used to describe the time course of a parent 

chemical or metabolite in an animal system. There are two types of pharmacokinetic models: data-based 

and physiologically-based. A data-based model divides the animal system into a series of compartments 

which, in general, do not represent real, identifiable anatomic regions of the body whereby the 

physiologically-based model compartments represent real anatomic regions of the body. 

Physiologically Based Pharmacodynamic (PBPD) Model—A type of physiologically-based doseresponse model which quantitatively describes the relationship between target tissue dose and toxic end 
HEXACHLOROBENZENE 350 

10. GLOSSARY 

points. These models advance the importance of physiologically based models in that they clearly 

describe the biological effect (response) produced by the system following exposure to an exogenous 

substance. 

Physiologically Based Pharmacokinetic (PBPK) Model—Comprised of a series of compartments 

representing organs or tissue groups with realistic weights and blood flows. These models require a 

variety of physiological information: tissue volumes, blood flow rates to tissues, cardiac output, alveolar 

ventilation rates and, possibly membrane permeabilities. The models also utilize biochemical information 

such as air/blood partition coefficients, and metabolic parameters. PBPK models are also called 

biologically based tissue dosimetry models. 

Prevalence—The number of cases of a disease or condition in a population at one point in time. 

Prospective Study—A type of cohort study in which the pertinent observations are made on events 

occurring after the start of the study. A group is followed over time. 

q1*—The upper-bound estimate of the low-dose slope of the dose-response curve as determined by the 

multistage procedure. The q1* can be used to calculate an estimate of carcinogenic potency, the 

incremental excess cancer risk per unit of exposure (usually μg/L for water, mg/kg/day for food, and 

μg/m3

 for air). 

Recommended Exposure Limit (REL)—A National Institute for Occupational Safety and Health 

(NIOSH) time-weighted average (TWA) concentrations for up to a 10-hour workday during a 40-hour 

workweek. 

Reference Concentration (RfC)—An estimate (with uncertainty spanning perhaps an order of 

magnitude) of a continuous inhalation exposure to the human population (including sensitive subgroups) 

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

The inhalation reference concentration is for continuous inhalation exposures and is appropriately 

expressed in units of mg/m3 or ppm. 

Reference Dose (RfD)—An estimate (with uncertainty spanning perhaps an order of magnitude) of the 

daily exposure of the human population to a potential hazard that is likely to be without risk of deleterious 

effects during a lifetime. The RfD is operationally derived from the no-observed-adverse-effect level 

(NOAEL-from animal and human studies) by a consistent application of uncertainty factors that reflect 

various types of data used to estimate RfDs and an additional modifying factor, which is based on a 

professional judgment of the entire database on the chemical. The RfDs are not applicable to 

nonthreshold effects such as cancer. 

Reportable Quantity (RQ)—The quantity of a hazardous substance that is considered reportable under 

the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA). Reportable 

quantities are (1) 1 pound or greater or (2) for selected substances, an amount established by regulation 

either under CERCLA or under Section 311 of the Clean Water Act. Quantities are measured over a 

24-hour period. 

Reproductive Toxicity—The occurrence of adverse effects on the reproductive system that may result 

from exposure to a chemical. The toxicity may be directed to the reproductive organs and/or the related 

endocrine system. The manifestation of such toxicity may be noted as alterations in sexual behavior, 

fertility, pregnancy outcomes, or modifications in other functions that are dependent on the integrity of 

this system. 
HEXACHLOROBENZENE 351 

10. GLOSSARY 

Retrospective Study—A type of cohort study based on a group of persons known to have been exposed 

at some time in the past. Data are collected from routinely recorded events, up to the time the study is 

undertaken. Retrospective studies are limited to causal factors that can be ascertained from existing 

records and/or examining survivors of the cohort. 

Risk—The possibility or chance that some adverse effect will result from a given exposure to a chemical. 

Risk Factor—An aspect of personal behavior or lifestyle, an environmental exposure, or an inborn or 

inherited characteristic, that is associated with an increased occurrence of disease or other health-related 

event or condition. 

Risk Ratio—The ratio of the risk among persons with specific risk factors compared to the risk among 

persons without risk factors. A risk ratio greater than 1 indicates greater risk of disease in the exposed 

group compared to the unexposed. 

Short-Term Exposure Limit (STEL)—The American Conference of Governmental Industrial 

Hygienists (ACGIH) maximum concentration to which workers can be exposed for up to 15 min 

continually. No more than four excursions are allowed per day, and there must be at least 60 min 

between exposure periods. The daily Threshold Limit Value - Time Weighted Average (TLV-TWA) may 

not be exceeded.

 Mortality Ratio (SMR)—A ratio of the observed number of deaths and the expected 

number of deaths in a specific standard population. 

Target Organ Toxicity—This term covers a broad range of adverse effects on target organs or 

physiological systems (e.g., renal, cardiovascular) extending from those arising through a single limited 

exposure to those assumed over a lifetime of exposure to a chemical. 

Teratogen—A chemical that causes structural defects that affect the development of an organism. 

Threshold Limit Value (TLV)—An American Conference of Governmental Industrial Hygienists 

(ACGIH) concentration of a substance to which most workers can be exposed without adverse effect. 

The TLV may be expressed as a Time Weighted Average (TWA), as a Short-Term Exposure Limit 

(STEL), or as a ceiling limit (CL). 

Time-Weighted Average (TWA)—An allowable exposure concentration averaged over a normal 8-hour 

workday or 40-hour workweek. 

Toxic Dose(50) (TD50)—A calculated dose of a chemical, introduced by a route other than inhalation, 

which is expected to cause a specific toxic effect in 50% of a defined experimental animal population. 

Toxicokinetic—The study of the absorption, distribution and elimination of toxic compounds in the 

living organism. 

Uncertainty Factor (UF)—A factor used in operationally deriving the Minimal Risk Level (MRL) or 

Reference Dose (RfD) or Reference Concentration (RfC) from experimental data. UFs are intended to 

account for (1) the variation in sensitivity among the members of the human population, (2) the 

uncertainty in extrapolating animal data to the case of human, (3) the uncertainty in extrapolating from 

data obtained in a study that is of less than lifetime exposure, and (4) the uncertainty in using lowestobserved-adverse-effect level (LOAEL) data rather than no-observed-adverse-effect level (NOAEL) data. 
HEXACHLOROBENZENE 352 

10. GLOSSARY 

A default for each individual UF is 10; if complete certainty in data exists, a value of one can be used; 

however a reduced UF of three may be used on a case-by-case basis, three being the approximate 

logarithmic average of 10 and 1. 

Xenobiotic—Any chemical that is foreign to the biological system. 

HEXACHLOROBENZENE A-1 

APPENDIX A
 

ATSDR MINIMAL RISK LEVEL AND WORKSHEETS
 

The Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) [42 U.S.C. 

9601 et seq.], as amended by the Superfund Amendments and Reauthorization Act (SARA) [Pub. L. 

99–499], requires that the Agency for Toxic Substances and Disease Registry (ATSDR) develop jointly 

with the U.S. Environmental Protection Agency (EPA), in order of priority, a list of hazardous substances 

most commonly found at facilities on the CERCLA National Priorities List (NPL); prepare toxicological 

profiles for each substance included on the priority list of hazardous substances; and assure the initiation 

of a research program to fill identified data needs associated with the substances. 

The toxicological profiles include an examination, summary, and interpretation of available toxicological 

information and epidemiologic evaluations of a hazardous substance. During the development of 

toxicological profiles, Minimal Risk Levels (MRLs) are derived when reliable and sufficient data exist to 

identify the target organ(s) of effect or the most sensitive health effect(s) for a specific duration for a 

given route of exposure. An MRL is an estimate of the daily human exposure to a hazardous substance 

that is likely to be without appreciable risk of adverse noncancer health effects over a specified duration 

of exposure. MRLs are based on noncancer health effects only and are not based on a consideration of 

cancer effects. These substance-specific estimates, which are intended to serve as screening levels, are 

used by ATSDR health assessors to identify contaminants and potential health effects that may be of 

concern at hazardous waste sites. It is important to note that MRLs are not intended to define clean-up or 

action levels. 

MRLs are derived for hazardous substances using the no-observed-adverse-effect level/uncertainty factor 

approach. They are below levels that might cause adverse health effects in the people most sensitive to 

such chemical-induced effects. MRLs are derived for acute (1–14 days), intermediate (15–364 days), and 

chronic (365 days and longer) durations and for the oral and inhalation routes of exposure. Currently, 

MRLs for the dermal route of exposure are not derived because ATSDR has not yet identified a method 

suitable for this route of exposure. MRLs are generally based on the most sensitive chemical-induced end 

point considered to be of relevance to humans. Serious health effects (such as irreparable damage to the 

liver or kidneys, or birth defects) are not used as a basis for establishing MRLs. Exposure to a level 

above the MRL does not mean that adverse health effects will occur. 
HEXACHLOROBENZENE A-2 

APPENDIX A 

MRLs are intended only to serve as a screening tool to help public health professionals decide where to 

look more closely. They may also be viewed as a mechanism to identify those hazardous waste sites that 

are not expected to cause adverse health effects. Most MRLs contain a degree of uncertainty because of 

the lack of precise toxicological information on the people who might be most sensitive (e.g., infants, 

elderly, nutritionally or immunologically compromised) to the effects of hazardous substances. ATSDR 

uses a conservative (i.e., protective) approach to address this uncertainty consistent with the public health 

principle of prevention. Although human data are preferred, MRLs often must be based on animal studies 

because relevant human studies are lacking. In the absence of evidence to the contrary, ATSDR assumes 

that humans are more sensitive to the effects of hazardous substance than animals and that certain persons 

may be particularly sensitive. Thus, the resulting MRL may be as much as a hundredfold below levels 

that have been shown to be nontoxic in laboratory animals. 

Proposed MRLs undergo a rigorous review process: Health Effects/MRL Workgroup reviews within the 

Division of Toxicology, expert panel peer reviews, and agencywide MRL Workgroup reviews, with 

participation from other federal agencies and comments from the public. They are subject to change as 

new information becomes available concomitant with updating the toxicological profiles. Thus, MRLs in 

the most recent toxicological profiles supersede previously published levels. For additional information 

regarding MRLs, please contact the Division of Toxicology, Agency for Toxic Substances and Disease 

Registry, 1600 Clifton Road, Mailstop E-29, Atlanta, Georgia 30333. 
HEXACHLOROBENZENE A-3 

APPENDIX A 

MINIMAL RISK LEVEL (MRL) WORKSHEET 

Chemical name: Hexachlorobenzene
 

CAS number: 118-74-1
 

Date: May 2002
 

Profile status: Draft 3 Post-Public Comment
 

Route: [ ] Inhalation [X] Oral
 

Duration: [X] Acute [ ] Intermediate [ ] Chronic
 

Key to figure: 23
 

Species: Rat
 

MRL: 0.008 [X] mg/kg/day [ ] ppm [ ] mg/m3
 

(LOAEL = 2.5 mg/kg/day; total uncertainty factors = 300)
 

Reference: Goldey ES and Taylor DH. 1992. Developmental neurotoxicity following premating 

maternal exposure to hexachlorobenzene in rats. Neurotoxicol and Teratol, 14: 15–21. 

Experimental design (human study details or strain, number of animals per exposure/control group, sex, 

dose administration details): Groups of 30 virgin female Sprague-Dawley rats were dosed by gastric 

intubation for 4 days with 0, 2.5, or 25 mg/kg/day hexachlorobenzene to achieve a total dose of 0, 10, or 

100 mg/kg for the 4-day period. Dosing was completed 2 weeks before breeding. The developmental 

neurotoxicity of hexachlorobenzene was assessed using a battery of behavioral tests. Negative geotaxic 

response was assessed in two male and two female pups from each litter on postnatal day (PND) 6, 8, and 

10. Olfactory discrimination/homing was assessed in two male and two female pups from each litter on 

PND 9, 10, and 11. This test simultaneously measures sensory discrimination, motivation, and locomotor 

ability. The development of exploration and locomotion was assessed in whole litters between PND 

15 and 20. Acoustic startle response (ASR) was assessed on PND 23 and 90. Visual discrimination 

learning, as measured in the water-filled T-maze, was assessed in offspring on PND 40. Motor activity in 

mature offspring (PND 60) was measured in an open area. These adult animals were again tested for 

exploratory activity on PND 100. 

Effects noted in study and corresponding doses: Hexachlorobenzene affected multiple pathways 

throughout the developing nervous system, manifested as slight hyperactivity, at a LOAEL of 

2.5 mg/kg/day. The offspring rats showed faster response times in negative geotaxis and olfactory 

discrimination/homing tests at the 2.5 or 25 mg/kg/day maternal dose level. Offspring exposed to 

maternal doses of 2.5 or 25 mg/kg/day showed either significantly increased exploratory behavior, slight 

hyperactivity, or both during the early life (19–20 days of age). Hexachlorobenzene-exposed offspring at 

the 25 mg/kg/day dose level exhibited significantly decreased ASR (23-day-old pups). When rats were 

tested later as adults (90 days old), response amplitude was significantly elevated in males in both groups 

exposed in utero to 2.5 and 25 mg/kg/day, compared to controls. Maternal exposure of rats to 

hexachlorobenzene did not result in any significant changes in learning ability, locomotor activity 

(60-day-old offspring), or exploratory activity (100-day-old offspring). 

Dose end point used for MRL derivation: 

[ ] NOAEL [X ] LOAEL 

2.5 mg/kg/day; hyperactivity in offspring. 
HEXACHLOROBENZENE A-4 

APPENDIX A 

Uncertainty factors used in MRL derivation: 

[ ] 1 [X] 3 [ ] 10 (for use of a minimal LOAEL)
 

[ ] 1 [ ] 3 [X] 10 (for extrapolation from rats to humans)
 

[ ] 1 [ ] 3 [X] 10 (for human variability)
 

Total uncertainty factors: 3 x 10 x 10 = 300
 

Was a conversion factor used from ppm in food or water to a mg/body weight dose?
 

No.
 

If an inhalation study in animals, list conversion factors used in determining human equivalent dose: 

Not applicable. 

Was a conversion used from intermittent to continuous exposure? 

No. 

Other additional studies or pertinent information that lend support to this MRL: Review of the located 

human and animal oral acute toxicity data for hexachlorobenzene indicate that the 4-day developmental 

study by Goldey and Taylor (1992) provides the most appropriate data for use deriving oral acute MRL 

for hexachlorobenzene. The LOAEL of 2.5 mg/kg/day is the most refined LOAEL for the acute toxicity 

of hexachlorobenzene. 

Adverse neurological signs and symptoms have also been observed in human offspring of maternally 

exposed to hexachlorobenzene. Children of mothers who had eaten hexachlorobenzene-contaminated 

bread (estimated hexachlorobenzene intake of 0.05–0.2 g/day) in Turkey exhibited muscle weakness, 

pinched facies, cogwheel rigidity, and sensory shading. Hexachlorobenzene was detected in the breast 

milk of the mothers, indicating lactational transfer (Cam and Nigogosyan 1963; Peters et al. 1982, 1987). 

Since hexachlorobenzene crosses the placenta and accumulates in fetal tissues in several animal species 

including the rat (Cripps 1990; Villeneuve and Hierlihy 1975), rabbit (Villeneuve et al. 1974a), and 

mouse (Courtney and Andrews 1985; Courtney et al. 1979), it is likely that the human offspring were also 

exposed to hexachlorobenzene during gestation. Development of neurotoxic signs has also been reported 

in other neonatal animals. These signs included convulsions, tremors, and progressive weakness in rats 

(Cripps 1990). Oral hexachlorobenzene has also been shown to interfere with the function of the nervous 

system in adult animals, inducing tremors, ataxia, and paralysis in unspecified strain of rats (Ockner and 

Schmid 1961); clonic convulsions, tremors, hyper-excitability, reversible muscle fasciculations, and 

lethargy in adult Wistar rats (Kennedy and Wigfield 1990; Koss et al. 1978; Nikolaev et al. 1986); mild 

reduction in conduction velocity of sciatic nerve, denervation, fibrillations, and chronic repetitive 

discharges in adult Sprague-Dawley rats (Sufit et al. 1986); tremor and hyperexcitability in adult Sherman 

rats (Kimbrough and Linder 1974); dysrhythmic electroencephalogram in adult Beagle dogs (Sundlof et 

al. 1981); tremor in adult C57B1/6J mice (Hahn et al. 1988); severe tremors and muscular weakness in 

adult Rhesus monkeys (Knauf and Hobson 1979); and tremors, panting, and unsteady gait in adult SPF 

pigs (Den Tonkelaar et al. 1978). 

Agency Contact (Chemical Manager): Dr. Jessilynn Taylor 
HEXACHLOROBENZENE A-5 

APPENDIX A 

MINIMAL RISK LEVEL WORKSHEET 

Chemical name: Hexachlorobenzene 

CAS number: 118-74-1 

Date: May 2002 

Profile status: Draft 3 Post-Public Comment 

Route: [ ] Inhalation [X ] Oral 

Duration: [ ] Acute [X] Intermediate [ ] Chronic 

Key to figure: 129 

Species: Monkey 

MRL: 0.0001 [X] mg/kg/day [ ] ppm [ ] mg/m3 

(LOAEL = 0.01 mg/kg/day; total uncertainty factors = 90) 

References: Bourque AC, Singh A, Lakhanpal N, et al. 1995. Ultrastructural changes in ovarian 

follicles of monkeys administered hexachlorobenzene. Am J Vet Res 56: 1673–1677. 

Jarrell JF, MacMahon A, Villeneuve D, et al. 1993. Hexachlorobenzene toxicity in the monkey 

primordial germ cell without induced porphyria. Reproductive Toxicology 7:41–47. 

Experimental design (human study details or strain, number of animals per exposure/control group, sex, 

dose administration details): In the Jarrell et al. (1993) study, groups of four female Cynomolgus 

monkeys were administered 0, 0.1, 1, or 10 mg/kg/day hexachlorobenzene in gelatin capsules for 90 days. 

No systemic toxicity was noted in any of the animals. After treatment, the animals were sacrificed and 

one ovary was removed from each animal from each dose group (including controls) and examined by 

transmission electron microscopy for alterations to surface epithelium. A cycle of in vitro fertilization 

with oocytes removed from exposed females during the menstrual cycle was performed to evaluate 

fertility. Induction of ovarian hyperstimulation (performed with human menopausal gonadotropin) was 

conducted to evaluate oocyte function. In the follow-up Bourque et al. (1995) study, groups of four 

female Cynomolgus monkeys were administered 0, 0.01, 0.1, 1, or 10 mg/kg/day hexachlorobenzene in 

gelatin capsules for 90 days. Monkeys were then given a preparation containing follicle-stimulating and 

luteinizing hormones on days 2–7 of the next menstrual cycle to stimulate follicle development, and 

human chorionic gonadotropin was administered on day 8 of the cycle. Oophorectomy was performed on 

day 10 of the cycle via laparotomy. The ovary was examined by transmission electron microscopy. 

Effects noted in study and corresponding doses: In the Jarrell et al. (1993) study, in vitro fertilization and 

ovarian hyperstimulation in terms of percent fertilization, estradiol response to gonadotropin, follicular 

development, oocyte recovery rates and maturation, and early embryo development were not significantly 

different from control animals. Hexachlorobenzene treatment caused a decrease in the total number of 

oocytes and primordial follicles. Ultrastructural changes in ovarian epithelium included a decrease in 

nuclear membrane distinction, an increase in density and granularity of oocyte nuclei, an increase in 

vacuoles, aggregated lysosomes in ooplasm of follicular cells, and pyknotic granulosa cells. The ooplasm 

of some follicles was necrotic, and some follicles had mild to moderate degenerative changes. These 

changes were observed in all exposed animals; the severity of symptoms increased in a dose-dependent 

manner. The follow-up study by Bourque et al. (1995) extended the observation of ultrastructural effects 

in the ovary to 0.01 mg/kg/day. At this dose, mitochondria in developing follicles were condensed and 

deformed. At higher doses, the mitochondria were progressively more damaged and additional changes, 

such as indentation of nuclear membranes and abnormal accumulation of lipid in the cytoplasm of 

follicular cells, were noted. 
HEXACHLOROBENZENE A-6 

APPENDIX A 

Dose end point used for MRL derivation: 

[ ] NOAEL [X ] LOAEL 

0.01 mg/kg/day; degenerative lesions in ovarian follicles 

Uncertainty factors used in MRL derivation: 

[ ] 1 [X] 3 [ ] 10 (for use of a minimal LOAEL)
 

[ ] 1 [X] 3 [ ] 10 (for extrapolation from monkeys to humans)
 

[ ] 1 [ ] 3 [X] 10 (for human variability)
 

Total uncertainty factors: 3 x 3 x 10 = 90
 

Was a conversion factor used from ppm in food or water to a mg/body weight dose?
 

No.
 

If an inhalation study in animals, list conversion factors used in determining human equivalent dose: 

Not applicable. 

Was a conversion used from intermittent to continuous exposure? 

No. 

Other additional studies or pertinent information that lend support to this MRL: Review of the located 

human and animal oral intermediate toxicity data for hexachlorobenzene indicate that the 90-day studies 

by Bourque et al. (1995) and Jarrell et al. (1993) provide the most appropriate data for use in deriving an 

oral intermediate MRL for hexachlorobenzene. 

In humans, 15 fetal deaths (comprising 13 miscarriages and 2 stillbirths) and 173 live births occurred in 

42 females in a 4-year period (1977–1981) several years after the widespread accidental ingestion of 

hexachlorobenzene-treated seed grain in Turkey (Peters et al. 1982, 1987). These mothers also had 

0.51 ppm hexachlorobenzene in their breast milk as compared to 0.07 ppm in unexposed controls 

(Gocmen et al. 1989). It has also been demonstrated that hexachlorobenzene in human milk can readily 

cross to the child during lactation and accumulate in the offspring (Ando et al. 1985; Weisenberg 1986; 

Weisenberg et al. 1985). 

Animal studies provide additional evidence that hexachlorobenzene is toxic to the mammalian ovary and 

may interfere with mechanisms regulating ovarian steroidogenesis. Female Cynomolgus monkeys 

exhibited a dose-dependent decrease in serum progesterone levels during the luteal phase of the menstrual 

cycle when administered hexachlorobenzene doses of $0.1 (0.1, 1, 10) mg/kg/day as capsules for 90 days; 

the decrease in levels of progesterone was not observed during the follicular and periovulatory phases. 

Lengthening of the menstrual cycle was also observed, as well as dose-dependent ultrastructural changes 

in surface epithelium of the ovary (indicative of cellular degeneration) and changes in ovary surface 

epithelial cell shape (length to width ratio), in all treatment groups, the severity of which was increased in 

a dose-dependent manner (Foster et al. 1992a; Sims et al. 1991). Increased serum progesterone levels and 

elevated ovarian weights were observed in superovulated female Sprague-Dawley rats orally administered 

$1 mg/kg/day hexachlorobenzene by gavage (in corn oil) for 21 days (Foster et al. 1992b). Serum levels 

of estradiol and progesterone of female Sprague-Dawley rats receiving daily doses of 50 mg/kg/day hexachlorobenzene by gavage in corn oil for 5 days were not significantly affected, although super-ovulated 

rats (dosed with pregnant mare serum gonadotropin and human chorionic gonadotropin) exposed to hexachlorobenzene in this study exhibited significant elevation of serum levels of progesterone (Foster et al. 

1993). In a subsequent study with ovariectomized female Sprague-Dawley rats administered daily oral 
HEXACHLOROBENZENE A-7 

APPENDIX A 

doses of 1, 10, or 100 mg/kg/day hexachlorobenzene in corn oil by gavage for 30 days, circulating levels 

of corticosterone levels were reduced by 25 and 51% at the 1 and 10 mg/kg/day hexachlorobenzene dose 

levels, respectively. Circulating cortisol levels were also significantly reduced (p<0.05). Hexachlorobenzene treatment had no effect on the levels of circulating aldosterone and progesterone levels, or on 

absolute and relative weights of the adrenal glands. The investigators concluded that hexachlorobenzene 

exposure induces alterations in steroidogenesis of cells of the adrenal cortex inner zone (Foster et al. 

1995b). 

A study conducted with female Rhesus monkeys given gavage doses of 8, 32, 64, or 128 mg/kg/day hexachlorobenzene in methylcellulose for 60 days found degenerative changes of the ovarian follicle, stroma, 

and germinal epithelium at dose levels of 64 mg/kg/day (Iatropoulos et al. 1976). In another monkey 

study, adult female Rhesus monkeys given oral doses of 8, 16, 32, 64, or 128 mg/kg/day hexachlorobenzene for 60 days showed significantly depressed (29%, p<0.01) whole serum cholesterol levels in 

weeks 3, 5, and 8. On day 60, depressed serum potassium and elevated SGOT were seen at the 

128 mg/kg/day dose level. The authors suggested that the changes in potassium and cholesterol levels 

may be due to liver histopathology. The authors suggested that the changes in potassium levels may be 

due to unusual steroidogenic activity associated with changes in ovarian morphology (Knauf and Hobson 

1979). 

Agency Contact (Chemical Manager): Dr. Jessilynn Taylor 
HEXACHLOROBENZENE A-8 

APPENDIX A 

MINIMAL RISK LEVEL WORKSHEET 

Chemical name: Hexachlorobenzene 

CAS number: 118-74-1 

Date: May 2002 

Profile status: Draft 3 Post-Public Comment 

Route: [ ] Inhalation [X] Oral 

Duration: [ ] Acute [ ] Intermediate [X] Chronic 

Key to figure: 162 

Species: Rat 

MRL: 0.00005 [X] mg/kg/day [ ] ppm [ ] mg/m3 

(LOAEL = 0.016 mg/kg/day; total uncertainty factors = 300) 

Reference: Arnold DL, Moodie CA, Charbonneau SM, Grice HC, et al. 1985. Long-term toxicity of 

hexachlorobenzene in the rat and the effect of dietary vitamin A. Food Chem Toxicol, 23(9): 779–793. 

Experimental design (human study details or strain, number of animals per exposure/control group, sex, 

dose administration details): Groups of Sprague-Dawley rats (40 per sex) of the F1 generation were 

exposed to dietary hexachlorobenzene at 0, 0.32, 1.6, 8, or 40 ppm (approximate doses of 0, 0.016, 0.08, 

0.4, or 2 mg/kg/day) from weaning for life (130 weeks). The groups of rats were also exposed in utero to 

hexachlorobenzene since variable numbers (40–66 per sex) of the F0 generation rats, from which the F1 

generation rats used in this study were randomly selected, were exposed to corresponding doses 3 months 

prior to mating and weaning of the F1 rats. The F0 rats were killed and examined at weaning of the F1 

rats. The F1 rats in this study were sacrificed after the animals had been on test for 130 weeks. A total of 

35 tissues and organs (including brain, heart, liver, extrahepatic bile duct, lungs, spleen, pancreas, small 

intestine, adrenals, kidneys, bladder, ovaries, uterus, skin, pituitary, thyroid, parathyroid, thymus, 

prostate, testes, and bone) were histopathologically examined. 

Effects noted in study and corresponding doses: Significant dose-response trends were observed in both 

sexes for hepatic basophilic chromogenesis at $0.4 mg/kg/day, and in males for peribiliary lymphocytosis 

and fibrosis at or greater than the lowest dose tested (0.016 mg/kg/day). Chronic nephrosis, severe in 

males, and reduced pup viability were observed at 2 mg/kg/day. Tumors were also increased at 2 

mg/kg/day, including neoplastic liver nodules in females, parathyroid adenoma in males, and adrenal 

pheochromocytoma in both males and females. No treatment related effects in the rat offspring were 

observed with respect to feed consumption or body weight. 

For derivation of the MRL, the increased incidences of peribiliary lymphocytosis and fibrosis in treated 

males were considered to represent a minimal effect. These are common spontaneous lesions in aging 

rats and occurred in approximately 30% of controls in this study. For peribiliary fibrosis, incidence was 

increased in all treated groups (statistically significant in the 0.016 and 2 mg/kg/day groups), but there 

was no clear evidence of a dose-response (13/48, 23/48, 21/48, 21/49, and 23/49 in the control, 0.016, 

0.08, 0.4, and 2 mg/kg/day groups, respectively). For peribiliary lymphocytosis, the incidence was 

increased in all treated groups (statistically significant in the 0.016, 0.08, and 2 mg/kg/day groups), and 

while the trend with dose was not very impressive, it was statistically significant (16/48, 27/48, 26/48, 

21/49, and 32/49, respectively). Incidences of these lesions in the control and treated females were 

similar to the control males (ranging from 6/49 to 14/49), suggesting that the incidence levels in control 

males were not unusually low. Overall, these findings suggest that hexachlorobenzene produced a 
HEXACHLOROBENZENE A-9 

APPENDIX A 

minimal hepatic effect in male rats at the lowest doses administered by increasing the incidence of agerelated hepatic lesions. 

Dose end point used for MRL derivation: 

[ ] NOAEL [X] LOAEL 

0.016 mg/kg/day; peribiliary lymphocytosis and fibrosis of the liver. 

Uncertainty factors used in MRL derivation: 

[ ] 1 [X ] 3 [] 10 (for use of a minimal LOAEL)
 

[ ] 1 [ ] 3 [X] 10 (for extrapolation from rats to humans)
 

[ ] 1 [ ] 3 [X] 10 (for human variability)
 

Total uncertainty factors: 3 x 10 x 10 = 300
 

Was a conversion factor used from ppm in food or water to a mg/body weight dose?
 

Yes. Dose = 0.32 ppm or 0.32 mg hexachlorobenzene/kg feed. The average rat consumes 0.05 kg of feed
 

per kg body weight per day (EPA 1986 reference food factor for rats). Thus, the average daily dose of
 

hexachlorobenzene is estimated to be 0.32 mg/kg x 0.05 kg feed/kg body weight/day = 0.016 mg/kg/day.
 

If an inhalation study in animals, list conversion factors used in determining human equivalent dose: 

Not applicable. 

Was a conversion used from intermittent to continuous exposure? 

No. 

Other additional studies or pertinent information that lend support to this MRL: The liver appears to be a 

target organ of hexachlorobenzene; therefore, using hepatic effects (peribiliary lymphocytosis and 

fibrosis) to calculate the MRL is appropriate. Also, review of the located human and animal oral chronic 

toxicity data for hexachlorobenzene indicate that the 130-week rat study by Arnold et al. (1985) provide 

the most appropriate data to derive a chronic MRL because the study provides the most refined LOAEL 

for the characteristic chronic toxicity (hepatic effects) of hexachlorobenzene. 

Other studies in several animal species have demonstrated that the liver is the major target organ of 

hexachlorobenzene exposure. Typical signs included microscopic lesions, increased porphyrin levels, 

and interference with hepatic enzymes involved in the heme biosynthesis pathway. 

Oral exposure to hexachlorobenzene induced liver histopathology and altered liver histochemistry in the 

rats. The relative liver weights of male rats were increased by 46% while those of females were increased 

by 23% in a study in which both sexes of Sprague-Dawley rats were given oral hexachlorobenzene doses 

of 27.5 mg/kg/day for 4 weeks (Richter et al. 1981). Cytoplasmic vacuolation, anisokaryosis, and 

pyknotic hepatocytes were seen in the liver of female Sprague-Dawley rats administered single gavage 

doses of 400 or 600 mg/kg hexachlorobenzene in corn oil and observed for 14 days. Relative liver 

weights were increased by 16–18 and 13–18% in the 400 and 600 mg/kg dose group animals, 

respectively. Serum cholesterol levels were increased by 13–30 and 7–31% in the 400 and 600 mg/kg 

dose group animals, respectively. No changes in serum sodium, potassium, inorganic phosphorus, total 

bilirubin, alkaline phosphatase, aspartate aminotransferase, total protein, calcium, glucose, and lactate 

dehydrogenase content were found. Liver microsomal aminopyrine demethylase, aniline hydroxylase, 

and ethoxyresorufin deethylase (EROD) activities were unchanged by hexachlorobenzene exposure 

(Lecavalier et al. 1994). Liver weight was significantly increased by nearly 45% in animals treated with 
HEXACHLOROBENZENE A-10 

APPENDIX A 

1,000 mg/kg/day for 7 days. Liver porphyrin carboxylase activity was significantly decreased in animals 

receiving 1,000 mg/kg/day (Kleiman de Pisarev et al. 1990). In rats given 100 mg/kg/day of hexachlorobenzene for 60 days, the liver became enlarged, and liver degeneration and increased hepatic porphyrins 

were observed (Ockner and Schmid 1961). A 67% hepatomegaly was observed in male Wistar 

(WAG/MBL) rats given oral doses of 120 mg/kg/day hexachlorobenzene, 3 times a week for 4 weeks 

(van Raaij et al. 1993b). Similarly, an 81% liver enlargement was reported at a hexachlorobenzene dose 

of 10 mg/kg/day for a group of Fischer 344 rats in a 15-week study. Liver-to-body weight ratios also 

increased at dose levels of 10 and 25 mg/kg/day (Andrews et al. 1989). Similarly, significantly increased 

liver weights were observed in rats fed 100 mg/kg/day hexachlorobenzene in the diet for 5 days 

(Rajamanickam and Padmanaban 1974). 

Porphyria, as an index of hepatopathology, was also seen in several animal studies following oral 

exposure to hexachlorobenzene. Uroporphyria and hepatic porphyrin accumulation were reported in a 

13-week study in female Wistar rats administered 7.5 or 15 mg/kg/day hexachlorobenzene in the feed. 

Rats in the 7.5 and 15 mg/kg/day dose group exhibited elevated d-ALA synthase levels (94 and 483%, 

respectively). At the low dose (7.5 mg/kg/day), relative liver weight increased by 31%, and at the highest 

dose tested (15 mg/kg/day), relative liver weight increased by 103%. Hypertrophic hepatocytes with 

eosinophilic cytoplasm with thready basophilic structures, as well as inflammatory cell infiltrates, were 

observed in the livers of animals dosed with 15 mg/kg/day hexachlorobenzene. By the end of the study, 

uroporphyria in the 15 mg/kg/day dose group was 1,400% that of control animals. Liver accumulations 

of porphyrins were 1,054 and 15,104%, respectively, for the 7.5 and 15 mg/kg/day dose group animals 

compared to undosed controls. Liver retinoid and plasma retinol were decreased by 70 and 53%, 

respectively, at the 15 mg/kg/day dose level (den Besten et al. 1993). 

Female Sprague-Dawley rats exhibited elevated urinary and hepatic porphyrins following administration 

of 50 mg/kg/day hexachlorobenzene for 12 consecutive days (cumulative dose of 600 mg/kg), and 

10 days (2 weeks, 5 days/week, cumulative dose of 500 mg/kg), or 100 mg/kg/day for 5 consecutive days 

(cumulative dose of 500 mg/kg). The porphyria seen was reported to be similar in severity to that 

observed in female rats receiving a cumulative dose of 1,500 mg/kg over a 6-week period. Porphyria 

induced by a cumulative dose of 500 mg/kg (either protocol) persisted for more than 500–600 days after 

exposure (Krishnan et al. 1991). Similarly, female Wistar rats given hexachlorobenzene in the diet at a 

dose of 150 mg/kg/day for 107 days exhibited a 91% decrease in uroporphyrinogen decarboxylase 

activity and 2,888-fold increase in hepatic porphyrin concentration (Elder and Urquhart 1986). 

Oral exposure to hexachlorobenzene resulted in altered liver function and histology in: adult female 

Rhesus monkeys given oral doses of 8, 16, 32, 64, or 128 mg/kg/day hexachlorobenzene for 60 days 

(Knauf and Hobson 1979); female Wistar rats exposed to 50 mg/kg/day hexachlorobenzene by gavage for 

15 weeks (Koss et al. 1978, 1983); female Agus Wistar rats fed diets containing 5 mg/kg/day hexachlorobenzene (with 2% arachis oil) for 75–90 weeks (Smith and Cabral 1980); female Wistar rats exposed to 

5 or 50 mg/kg hexachlorobenzene for 56 days (Kennedy and Wigfield 1990); both sexes of Charles River 

rats fed hexachlorobenzene doses of 0.5, 2, 8, or 32 mg/kg/day (Kuiper-Goodman et al. 1977); monkeys 

given doses of 8 mg/kg/day for 60 days (Iatropoulos et al. 1976); and adult female Beagle dogs 

administered $25 mg/kg/day hexachlorobenzene in gelatin capsules for 21 days (Sundlof et al. 1981). 

Induction of liver microsomal enzymes, increased liver weight, and microscopic lesions were 

demonstrated at a hexachlorobenzene dose of 5 mg/kg/day, while centrilobular hypertrophy, elevated 

urinary coproporphyrinogen, and depressed glucose-6-phosphatase activity were observed at a lower dose 

level (0.5 mg/kg/day) in pigs treated for 90 days (Den Tonkelaar et al. 1978). 

Several animal studies also found increased induction of P-450 isozymes and other hepatic enzymes, 

usually accompanied by hepatic or uroporphyria, as an index of adverse effects in the liver (Adjarov et al. 

1982; Hahn et al. 1988, 1989; Kitchin and Brown 1989; Kleiman de Pisarev et al. 1995; Li et al. 1989; 

Linko et al. 1986; Lissner et al. 1975; Mehendale et al. 1975; Rajamanickam and Padmanaban 1974; 

Smith et al. 1985; Wada et al. 1968). 
HEXACHLOROBENZENE A-11 

APPENDIX A 

Agency Contact (Chemical Manager): Dr. Jessilynn Taylor 

HEXACHLOROBENZENE B-1 

APPENDIX B 

USER'S GUIDE 

Chapter 1 

Public Health Statement 

This chapter of the profile is a health effects summary written in non-technical language. Its intended 

audience is the general public especially people living in the vicinity of a hazardous waste site or 

chemical release. If the Public Health Statement were removed from the rest of the document, it would 

still communicate to the lay public essential information about the chemical. 

The major headings in the Public Health Statement are useful to find specific topics of concern. The 

topics are written in a question and answer format. The answer to each question includes a sentence that 

will direct the reader to chapters in the profile that will provide more information on the given topic. 

Chapter 2 

Relevance to Public Health 

This chapter provides a health effects summary based on evaluations of existing toxicologic, 

epidemiologic, and toxicokinetic information. This summary is designed to present interpretive, 

weight-of-evidence discussions for human health end points by addressing the following questions. 

1.	 What effects are known to occur in humans? 

2.	 What effects observed in animals are likely to be of concern to humans? 

3.	 What exposure conditions are likely to be of concern to humans, especially around hazardous 

waste sites? 

The chapter covers end points in the same order they appear within the Discussion of Health Effects by 

Route of Exposure section, by route (inhalation, oral, dermal) and within route by effect. Human data are 

presented first, then animal data. Both are organized by duration (acute, intermediate, chronic). In vitro 

data and data from parenteral routes (intramuscular, intravenous, subcutaneous, etc.) are also considered 

in this chapter. If data are located in the scientific literature, a table of genotoxicity information is 

included. 

The carcinogenic potential of the profiled substance is qualitatively evaluated, when appropriate, using 

existing toxicokinetic, genotoxic, and carcinogenic data. ATSDR does not currently assess cancer 

potency or perform cancer risk assessments. Minimal risk levels (MRLs) for noncancer end points (if 

derived) and the end points from which they were derived are indicated and discussed. 

Limitations to existing scientific literature that prevent a satisfactory evaluation of the relevance to public 

health are identified in the Chapter 3 Data Needs section. 
HEXACHLOROBENZENE B-2 

APPENDIX B 

Interpretation of Minimal Risk Levels 

Where sufficient toxicologic information is available, we have derived minimal risk levels (MRLs) for 

inhalation and oral routes of entry at each duration of exposure (acute, intermediate, and chronic). These 

MRLs are not meant to support regulatory action; but to acquaint health professionals with exposure 

levels at which adverse health effects are not expected to occur in humans. They should help physicians 

and public health officials determine the safety of a community living near a chemical emission, given the 

concentration of a contaminant in air or the estimated daily dose in water. MRLs are based largely on 

toxicological studies in animals and on reports of human occupational exposure. 

MRL users should be familiar with the toxicologic information on which the number is based. Chapter 2, 

"Relevance to Public Health," contains basic information known about the substance. Other sections such 

as Chapter 3 Section 3.9, "Interactions with Other Substances,” and Section 3.10, "Populations that are 

Unusually Susceptible" provide important supplemental information. 

MRL users should also understand the MRL derivation methodology. MRLs are derived using a 

modified version of the risk assessment methodology the Environmental Protection Agency (EPA) 

provides (Barnes and Dourson 1988) to determine reference doses for lifetime exposure (RfDs). 

To derive an MRL, ATSDR generally selects the most sensitive end point which, in its best judgement, 

represents the most sensitive human health effect for a given exposure route and duration. ATSDR 

cannot make this judgement or derive an MRL unless information (quantitative or qualitative) is available 

for all potential systemic, neurological, and developmental effects. If this information and reliable 

quantitative data on the chosen end point are available, ATSDR derives an MRL using the most sensitive 

species (when information from multiple species is available) with the highest NOAEL that does not 

exceed any adverse effect levels. When a NOAEL is not available, a lowest-observed-adverse-effect 

level (LOAEL) can be used to derive an MRL, and an uncertainty factor (UF) of 10 must be employed. 

Additional uncertainty factors of 10 must be used both for human variability to protect sensitive 

subpopulations (people who are most susceptible to the health effects caused by the substance) and for 

interspecies variability (extrapolation from animals to humans). In deriving an MRL, these individual 

uncertainty factors are multiplied together. The product is then divided into the inhalation concentration 

or oral dosage selected from the study. Uncertainty factors used in developing a substance-specific MRL 

are provided in the footnotes of the LSE Tables. 

Chapter 3 

Health Effects 

Tables and Figures for Levels of Significant Exposure (LSE) 

Tables (3-1, 3-2, and 3-3) and figures (3-1 and 3-2) are used to summarize health effects and illustrate 

graphically levels of exposure associated with those effects. These levels cover health effects observed at 

increasing dose concentrations and durations, differences in response by species, minimal risk levels 

(MRLs) to humans for noncancer end points, and EPA's estimated range associated with an upper- bound 

individual lifetime cancer risk of 1 in 10,000 to 1 in 10,000,000. Use the LSE tables and figures for a 

quick review of the health effects and to locate data for a specific exposure scenario. The LSE tables and 

figures should always be used in conjunction with the text. All entries in these tables and figures 

represent studies that provide reliable, quantitative estimates of No-Observed-Adverse-Effect Levels 

(NOAELs), Lowest-Observed-Adverse-Effect Levels (LOAELs), or Cancer Effect Levels (CELs). 
HEXACHLOROBENZENE B-3 

APPENDIX B 

The legends presented below demonstrate the application of these tables and figures. Representative 

examples of LSE Table 3-1 and Figure 3-1 are shown. The numbers in the left column of the legends 

correspond to the numbers in the example table and figure. 

LEGEND 

See LSE Table 3-1 

(1)	 Route of Exposure One of the first considerations when reviewing the toxicity of a substance using 

these tables and figures should be the relevant and appropriate route of exposure. When sufficient 

data exists, three LSE tables and two LSE figures are presented in the document. The three LSE 

tables present data on the three principal routes of exposure, i.e., inhalation, oral, and dermal (LSE 

Table 3-1, 3-2, and 3-3, respectively). LSE figures are limited to the inhalation (LSE Figure 3-1) 

and oral (LSE Figure 3-2) routes. Not all substances will have data on each route of exposure and 

will not therefore have all five of the tables and figures. 

(2)	 Exposure Period Three exposure periods - acute (less than 15 days), intermediate (15–364 days), 

and chronic (365 days or more) are presented within each relevant route of exposure. In this 

example, an inhalation study of intermediate exposure duration is reported. For quick reference to 

health effects occurring from a known length of exposure, locate the applicable exposure period 

within the LSE table and figure. 

(3)	 Health Effect The major categories of health effects included in LSE tables and figures are death, 

systemic, immunological, neurological, developmental, reproductive, and cancer. NOAELs and 

LOAELs can be reported in the tables and figures for all effects but cancer. Systemic effects are 

further defined in the "System" column of the LSE table (see key number 18). 

(4)	 Key to Figure Each key number in the LSE table links study information to one or more data points 

using the same key number in the corresponding LSE figure. In this example, the study represented 

by key number 18 has been used to derive a NOAEL and a Less Serious LOAEL (also see the 2 

"18r" data points in Figure 3-1). 

(5)	 Species The test species, whether animal or human, are identified in this column. Chapter 2, 

"Relevance to Public Health," covers the relevance of animal data to human toxicity and 

Section 3.4, "Toxicokinetics," contains any available information on comparative toxicokinetics. 

Although NOAELs and LOAELs are species specific, the levels are extrapolated to equivalent 

human doses to derive an MRL. 

(6)	 Exposure Frequency/Duration The duration of the study and the weekly and daily exposure 

regimen are provided in this column. This permits comparison of NOAELs and LOAELs from 

different studies. In this case (key number 18), rats were exposed to 1,1,2,2-tetrachloroethane via 

inhalation for 6 hours per day, 5 days per week, for 3 weeks. For a more complete review of the 

dosing regimen refer to the appropriate sections of the text or the original reference paper, i.e., 

Nitschke et al. 1981. 

(7)	 System This column further defines the systemic effects. These systems include: respiratory, 

cardiovascular, gastrointestinal, hematological, musculoskeletal, hepatic, renal, and dermal/ocular. 

"Other" refers to any systemic effect (e.g., a decrease in body weight) not covered in these systems. 

In the example of key number 18, 1 systemic effect (respiratory) was investigated. 
HEXACHLOROBENZENE B-4 

APPENDIX B 

(8)	 NOAEL A No-Observed-Adverse-Effect Level (NOAEL) is the highest exposure level at which no 

harmful effects were seen in the organ system studied. Key number 18 reports a NOAEL of 3 ppm 

for the respiratory system which was used to derive an intermediate exposure, inhalation MRL of 

0.005 ppm (see footnote "b"). 

(9)	 LOAEL A Lowest-Observed-Adverse-Effect Level (LOAEL) is the lowest dose used in the study 

that caused a harmful health effect. LOAELs have been classified into "Less Serious" and 

"Serious" effects. These distinctions help readers identify the levels of exposure at which adverse 

health effects first appear and the gradation of effects with increasing dose. A brief description of 

the specific end point used to quantify the adverse effect accompanies the LOAEL. The respiratory 

effect reported in key number 18 (hyperplasia) is a Less serious LOAEL of 10 ppm. MRLs are not 

derived from Serious LOAELs. 

(10)	 Reference The complete reference citation is given in Chapter 9 of the profile. 

(11)	 CEL A Cancer Effect Level (CEL) is the lowest exposure level associated with the onset of 

carcinogenesis in experimental or epidemiologic studies. CELs are always considered serious 

effects. The LSE tables and figures do not contain NOAELs for cancer, but the text may report 

doses not causing measurable cancer increases. 

(12)	 Footnotes Explanations of abbreviations or reference notes for data in the LSE tables are found in 

the footnotes. Footnote "b" indicates the NOAEL of 3 ppm in key number 18 was used to derive an 

MRL of 0.005 ppm. 

LEGEND 

See Figure 3-1 

LSE figures graphically illustrate the data presented in the corresponding LSE tables. Figures help the 

reader quickly compare health effects according to exposure concentrations for particular exposure 

periods. 

(13)	 Exposure Period The same exposure periods appear as in the LSE table. In this example, health 

effects observed within the intermediate and chronic exposure periods are illustrated. 

(14)	 Health Effect These are the categories of health effects for which reliable quantitative data exists. 

The same health effects appear in the LSE table. 

(15)	 Levels of Exposure concentrations or doses for each health effect in the LSE tables are graphically 

displayed in the LSE figures. Exposure concentration or dose is measured on the log scale "y" axis. 

Inhalation exposure is reported in mg/m3

 or ppm and oral exposure is reported in mg/kg/day. 

(16)	 NOAEL In this example, 18r NOAEL is the critical end point for which an intermediate inhalation 

exposure MRL is based. As you can see from the LSE figure key, the open-circle symbol indicates 

to a NOAEL for the test species-rat. The key number 18 corresponds to the entry in the LSE table. 

The dashed descending arrow indicates the extrapolation from the exposure level of 3 ppm (see 

entry 18 in the Table) to the MRL of 0.005 ppm (see footnote "b" in the LSE table). 

(17)	 CEL Key number 38r is 1 of 3 studies for which Cancer Effect Levels were derived. The diamond 

symbol refers to a Cancer Effect Level for the test species-mouse. The number 38 corresponds to 

the entry in the LSE table. 
HEXACHLOROBENZENE B-5 

APPENDIX B 

(18)	 Estimated Upper-Bound Human Cancer Risk Levels This is the range associated with the 

upper-bound for lifetime cancer risk of 1 in 10,000 to 1 in 10,000,000. These risk levels are 

derived from the EPA's Human Health Assessment Group's upper-bound estimates of the slope of 

the cancer dose response curve at low dose levels (q1*). 

(19)	 Key to LSE Figure The Key explains the abbreviations and symbols used in the figure. 
6

SAMPLE 

123

4

12 

Table 3-1. Levels of Significant Exposure to [Chemical x] – Inhalation

Exposure

Key to frequency/figurea Species duration 

6 INTERMEDIATE EXPOSURE

5 6 

6 Systemic 9 9

6 18 Rat	 13 wk

5 d/wk6 hr/d 

CHRONIC EXPOSURECancer38 Rat 18 mo

5 d/wk7 hr/d 

39 Rat	 89–104 wk

5 d/wk6 hr/d 

40 Mouse	 79–103 wk

5 d/wk6 hr/d 

NOAEL

System (ppm) 

7 8 

9 9

Resp 3b 

LOAEL (effect) 

Less serious (ppm) Serious (ppm) Reference

9 10 

9 9

10 (hyperplasia)	 Nitschke et al.

1981 

11

9

20

1010 

(CEL, multiple

organs)

(CEL, lung tumors,

nasal tumors)

(CEL, lung tumors,

hemangiosarcomas) 

Wong et al. 1982 NTP 1982 NTP 1982 

a The number corresponds to entries in Figure 3-1.b Used to derive an intermediate inhalation Minimal Risk Level (MRL) of 5 x 10-3 ppm; dose adjusted for intermittent exposure and divided by

an uncertainty factor of 100 (10 for extrapolation from animal to humans, 10 for human variability). 

6

HEXACHLOROBENZENE B-6 

APPENDIX B 
HEXACHLOROBENZENE B-7 

APPENDIX B 

HEXACHLOROBENZENE C-1 

APPENDIX C 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS 

ACOEM American College of Occupational and Environmental Medicine 

ACGIH American Conference of Governmental Industrial Hygienists 

ADI acceptable daily intake 

ADME absorption, distribution, metabolism, and excretion 

AED atomic emission detection 

AOEC Association of Occupational and Environmental Clinics 

AFID alkali flame ionization detector 

AFOSH Air Force Office of Safety and Health 

ALT alanine aminotransferase 

AML acute myeloid leukemia 

AOAC Association of Official Analytical Chemists 

AP alkaline phosphatase 

APHA American Public Health Association 

AST aspartate aminotranferase 

atm atmosphere 

ATSDR Agency for Toxic Substances and Disease Registry 

AWQC Ambient Water Quality Criteria 

BAT best available technology 

BCF bioconcentration factor 

BEI Biological Exposure Index 

BSC Board of Scientific Counselors 

C centigrade 

CAA Clean Air Act 

CAG Cancer Assessment Group of the U.S. Environmental Protection Agency 

CAS Chemical Abstract Services 

CDC Centers for Disease Control and Prevention 

CEL cancer effect level 

CELDS Computer-Environmental Legislative Data System 

CERCLA Comprehensive Environmental Response, Compensation, and Liability Act 

CFR Code of Federal Regulations 

Ci curie 

CI confidence interval 

CL ceiling limit value 

CLP Contract Laboratory Program 

cm centimeter 

CML chronic myeloid leukemia 

CPSC Consumer Products Safety Commission 

CWA Clean Water Act 

DHEW Department of Health, Education, and Welfare 

DHHS Department of Health and Human Services 

DNA deoxyribonucleic acid 

DOD Department of Defense 

DOE Department of Energy 

DOL Department of Labor 

DOT Department of Transportation 
HEXACHLOROBENZENE C-2 

APPENDIX C 

DOT/UN/ Department of Transportation/United Nations/

 NA/IMCO North America/International Maritime Dangerous Goods Code 

DWEL drinking water exposure level 

ECD electron capture detection 

ECG/EKG electrocardiogram 

EEG electroencephalogram 

EEGL Emergency Exposure Guidance Level 

EPA Environmental Protection Agency 

F Fahrenheit 

F1 first-filial generation 

FAO Food and Agricultural Organization of the United Nations 

FDA Food and Drug Administration 

FEMA Federal Emergency Management Agency 

FIFRA Federal Insecticide, Fungicide, and Rodenticide Act 

FPD flame photometric detection 

fpm feet per minute 

FR Federal Register 

FSH follicle stimulating hormone 

g gram 

GC gas chromatography 

gd gestational day 

GLC gas liquid chromatography 

GPC gel permeation chromatography 

HPLC high-performance liquid chromatography 

HRGC high resolution gas chromatography 

HSDB Hazardous Substance Data Bank 

IARC International Agency for Research on Cancer 

IDLH immediately dangerous to life and health 

ILO International Labor Organization 

IRIS Integrated Risk Information System 

Kd adsorption ratio 

kg kilogram 

Koc organic carbon partition coefficient 

Kow octanol-water partition coefficient 

L liter 

LC liquid chromatography 

LCLo lethal concentration, low 

LC50 lethal concentration, 50% kill 

LDLo lethal dose, low 

LD50 lethal dose, 50% kill 

LDH lactic dehydrogenase 

LH luteinizing hormone 

LT50 lethal time, 50% kill 

LOAEL lowest-observed-adverse-effect level 

LSE Levels of Significant Exposure 

m meter 

MA trans,trans-muconic acid 

MAL maximum allowable level 

mCi millicurie 

MCL maximum contaminant level 

MCLG maximum contaminant level goal 
HEXACHLOROBENZENE C-3 

APPENDIX C 

MFO mixed function oxidase 

mg milligram 

mL milliliter 

mm millimeter 

mmHg millimeters of mercury 

mmol millimole 

mppcf millions of particles per cubic foot 

MRL Minimal Risk Level 

MS mass spectrometry 

NAAQS National Ambient Air Quality Standard 

NAS National Academy of Science 

NATICH National Air Toxics Information Clearinghouse 

NATO North Atlantic Treaty Organization 

NCE normochromatic erythrocytes 

NCEH National Center for Environmental Health 

NCI National Cancer Institute 

ND not detected 

NFPA National Fire Protection Association 

ng nanogram 

NIEHS National Institute of Environmental Health Sciences 

NIOSH National Institute for Occupational Safety and Health 

NIOSHTIC NIOSH's Computerized Information Retrieval System 

NLM National Library of Medicine 

nm nanometer 

NHANES National Health and Nutrition Examination Survey 

nmol nanomole 

NOAEL no-observed-adverse-effect level 

NOES National Occupational Exposure Survey 

NOHS National Occupational Hazard Survey 

NPD nitrogen phosphorus detection 

NPDES National Pollutant Discharge Elimination System 

NPL National Priorities List 

NR not reported 

NRC National Research Council 

NS not specified 

NSPS New Source Performance Standards 

NTIS National Technical Information Service 

NTP National Toxicology Program 

ODW Office of Drinking Water, EPA 

OERR Office of Emergency and Remedial Response, EPA 

OHM/TADS Oil and Hazardous Materials/Technical Assistance Data System 

OPP Office of Pesticide Programs, EPA 

OPPTS Office of Prevention, Pesticides and Toxic Substances, EPA 

OPPT Office of Pollution Prevention and Toxics, EPA 

OR odds ratio 

OSHA Occupational Safety and Health Administration 

OSW Office of Solid Waste, EPA 

OW Office of Water 

OWRS Office of Water Regulations and Standards, EPA 

PAH polycyclic aromatic hydrocarbon 

PBPD physiologically based pharmacodynamic 
HEXACHLOROBENZENE C-4 

APPENDIX C 

PBPK physiologically based pharmacokinetic 

PCE polychromatic erythrocytes 

PEL permissible exposure limit 

PID photo ionization detector 

pg picogram 

pmol picomole 

PHS Public Health Service 

PMR proportionate mortality ratio 

ppb parts per billion 

ppm parts per million 

ppt parts per trillion 

PSNS pretreatment standards for new sources 

RBC red blood cell 

REL recommended exposure level/limit 

RfC reference concentration 

RfD reference dose 

RNA ribonucleic acid 

RTECS Registry of Toxic Effects of Chemical Substances 

RQ reportable quantity 

SARA Superfund Amendments and Reauthorization Act 

SCE sister chromatid exchange 

SGOT serum glutamic oxaloacetic transaminase 

SGPT serum glutamic pyruvic transaminase 

SIC standard industrial classification 

SIM selected ion monitoring 

SMCL secondary maximum contaminant level 

SMR standardized mortality ratio 

SNARL suggested no adverse response level 

SPEGL Short-Term Public Emergency Guidance Level 

STEL short term exposure limit 

STORET Storage and Retrieval 

TD50 toxic dose, 50% specific toxic effect 

TLV threshold limit value 

TOC total organic carbon 

TPQ threshold planning quantity 

TRI Toxics Release Inventory 

TSCA Toxic Substances Control Act 

TWA time-weighted average 

UF uncertainty factor 

U.S. United States 

USDA United States Department of Agriculture 

USGS United States Geological Survey 

VOC volatile organic compound 

WBC white blood cell 

WHO World Health Organization 

> greater than $ greater than or equal to 

= equal to 

< less than # less than or equal to 
HEXACHLOROBENZENE C-5 

APPENDIX C 

% percent α alpha β beta γ gamma δ delta 

μm micrometer 

μg microgram q1

* cancer slope factor 

– negative 

+ positive 

(+) weakly positive result 

(–) weakly negative result 

HEXACHLOROBENZENE D-1 

APPENDIX D 

INDEX 

adenocarcinoma . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 113, 178
 

adenoma . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 113, 178
 

adipose tissue . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 122, 132, 134, 141, 202, 236, 237, 241, 248, 249, 253, 254, 256, 265
 

adrenal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6, 15, 19, 88, 90, 106, 107, 111, 114, 116, 133, 151, 157, 161, 176, 178
 

adrenal gland . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6, 19, 88, 90, 114, 116, 157, 161
 

adrenal glands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 176 
 

adrenals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 132, 133
 

adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 197, 201, 211, 257
 

aerobic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 216-218
 

Ah receptor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167 
 

air . . . . . . . . . . . . . . . . . . . . . 2, 3, 5, 7, 9, 11, 15, 25, 26, 72, 74, 124-126, 142, 144, 159, 163, 170, 185, 191, 196, 197, 199, 201­

203, 206, 207, 210-213, 215, 216, 219-221, 235, 236, 239, 246, 247, 249, 257, 258, 265, 269, 271,
 

273, 275
 

ambient air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11, 74, 219, 220, 258
 

anaerobic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 217, 218, 246, 247
 

Arctic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 240 
 

BAF . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 214 
 

bass . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 213, 243
 

BCF . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 213-215, 248
 

bioaccumulation ............................................................................. 213, 247, 248
 

bioavailability ................................................................................... 247, 248
 

bioconcentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201, 213, 215, 247
 

bioconcentration factors (see BCF) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201 
 

biodegradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 189, 216-218, 246, 247
 

biomagnification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 213, 247, 248
 

biomarker . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 161, 162, 164, 183
 

birth weight . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74 
 

blood . . . . . . . . . . . . . . . . . . . . 3-5, 8, 12, 14, 15, 17, 25, 26, 72-75, 77, 79, 80, 86, 90, 93, 96-101, 103, 104, 109, 110, 113, 119,
 

121-123, 125-128, 130-139, 141, 143, 148-151, 158-160, 162, 163, 169, 171, 179-183, 235, 237, 

238, 249, 250, 253-255, 265, 266 

body weight effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 77, 94, 95 

breast cancer . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 113, 158, 178 

breast milk . . . . . . . . . . . . . . . . . . . 4, 8, 14, 73, 101, 103, 109, 122, 126, 133, 134, 136, 143, 146, 148, 151, 160, 161, 163, 181, 

182, 187, 239-241, 244, 250, 253, 254, 256, 265 

cancer . . . . . . . . . . . . . . . . . . . . . 4, 5, 9, 12, 15, 16, 24, 74, 112-115, 117, 156, 158, 159, 168, 174, 177, 178, 187, 190, 271, 274, 

275
 

carcinogen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5, 16, 112, 118, 178, 269, 273, 275
 

carcinogenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16, 23, 24, 156, 165, 178, 269, 275
 

carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 16, 75, 112, 178, 271-273
 

carcinoma . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12, 74, 112-114, 155, 188, 269
 

cardiovascular effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 78, 180 
 

catfish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 213, 214, 227, 234, 243
 

cholinesterase . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167 
 

Clean Water Act . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 273 
 

CYP1A1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167, 168
 

CYP1A2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 153, 167, 173
 

CYP2B1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 168 
 

death . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4, 12, 16, 23, 25, 75, 76, 95, 101, 108-111, 117, 167, 174, 176, 177, 180, 182
 

dechlorination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123, 139, 152, 217-219
 

degeneration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18, 21, 79, 81, 84, 86, 105-107
 

deiodinase . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 89 
 

deoxyribonucleic acid (see DNA) ................................................................... 120, 121
 

Department of Health and Human Services (see DHHS) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5 
 

dermal effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 91-93, 99, 108 
 
HEXACHLOROBENZENE D-2 

APPENDIX D 

developmental effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14, 74, 108, 111, 112, 117, 174, 176, 177, 180, 187
 

DHHS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5, 112 
 

DNA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 98, 117-121, 156, 162, 179
 

dye . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 196 
 

endocrine effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26, 72, 87, 88 
 

EROD . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 168 
 

estrogen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14, 88, 104-106, 156, 157
 

estrogen receptor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106 
 

FDA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9, 232, 233, 241, 250
 

FEDRIP . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 187, 251
 

fetus . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 122, 135, 136, 148, 158, 160
 

fibrosis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14, 20, 74, 78, 79, 84, 110, 112, 269
 

fish . . . . . . . . . . . . . . . . . . . . . 3, 87, 88, 96, 101, 109, 110, 125, 139, 182, 201, 202, 213, 214, 224-228, 232-235, 241-244, 247,
 

248, 260, 261, 263, 265, 274
 

fish consumption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 242, 243, 274
 

Flix . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12, 14, 15, 17, 25, 26, 73, 74, 119, 124, 125, 131, 135, 141, 160, 163, 182
 

flounder . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 224 
 

Food and Drug Administration (see FDA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9, 232 
 

fruits . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 262, 264
 

FSH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 103, 106
 

gastrointestinal effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 79 
 

general population . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11, 12, 15, 113, 161, 208, 234, 236-238, 242, 245, 247, 250
 

goiter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12, 72 
 

grass . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 209, 212, 214, 215
 

groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2, 3, 201, 202, 206, 212, 216, 242, 246, 247, 249, 250, 259, 260, 272, 274
 

half-life . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2, 124, 130, 135, 144, 148, 149, 162, 199, 201, 214, 216, 217, 219, 246
 

hemangiohepatoma . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 113 
 

hematological effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 79, 80 
 

Henry’s law . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 241 
 

hepatic effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12, 14, 19, 20, 25, 26, 81, 82, 84, 91, 96, 108, 113, 164, 174, 176
 

hepatocellular carcinomas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 114 
 

hepatoma . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 113, 114, 118, 178, 179
 

histopathology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18-20, 78, 84, 87, 88, 90, 93, 97-101, 106, 116, 176
 

hydrolysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 166, 216, 217, 247, 266
 

hydroxyl radical . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 155, 191
 

hyperplasia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14, 15, 84, 90, 93, 99, 100, 116, 178
 

hypothyroidism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6, 12, 72, 88, 89, 157, 161
 

IgA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 72, 93, 97 
 

IgG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 72, 92, 97-100, 111
 

immune system . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5, 14, 92, 96, 97, 99, 109, 177, 181
 

immunological effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12, 92, 96, 97, 99, 109, 174, 176, 177, 180, 186
 

insects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 209 
 

Integrated Risk Information System (see IRIS) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 112, 275
 

IRIS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 112, 269, 272, 275
 

kara yara . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13, 91, 108, 183
 

kidney . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5, 14, 16, 26, 77, 81, 84, 85, 87, 112, 113, 117, 123, 132, 133, 148, 149, 151, 167, 228
 

lake . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11, 193, 201, 208, 211, 213, 218-223, 225-227, 229, 231, 233, 234, 243, 247
 

LD50 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 75, 167 
 

leukemia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 113, 178
 

LH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106 
 

lichens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201, 202, 247, 248
 

liver . . . . . . . . . . . . . . . . . . . 4-6, 12-16, 19-21, 26, 74, 75, 77, 78, 81-84, 89, 90, 94, 96, 102, 105, 110-115, 117, 118, 122-124,
 

128-130, 132-136, 138, 139, 147-149, 151-153, 155, 157, 161, 162, 164-168, 172, 173, 176-178, 

181, 182, 226, 228, 231, 232, 236, 237, 244, 269
 

lung . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 73, 78, 93, 98, 99, 110, 111, 113, 117, 123, 132, 149, 151, 162
 

luteinizing hormone (see LH) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 87 
 

lymph . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 73, 94, 97-99, 111, 116, 119, 127, 132, 150, 151, 176
 

lymphatic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 122, 128, 132, 151, 248
 

lymphocytosis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14, 20, 84, 96, 110, 112, 269
 

lymphoreticular effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 72, 96, 117 
 

marine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201, 224, 231, 235
 
HEXACHLOROBENZENE D-3 

APPENDIX D 

marine mammals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201 
 

Mast cells . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 92, 93, 96 
 

menstrual . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18, 105, 157, 176, 179, 187
 

mercaptotetrachlorothioanisole . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123, 138
 

metaplasia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 115, 178
 

microproteinuria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12, 26 
 

milk . . . . . . . . . . . . . . . . . . . . 3, 4, 6, 8, 12, 14, 17, 73, 75, 96, 101, 103, 108, 109, 122, 126, 131, 133, 134, 136, 143, 146, 148,
 

151, 152, 159-161, 163, 169, 180-182, 187, 232, 237, 239-241, 244, 249, 250, 253, 254, 256, 261,
 

264, 265
 

Minimal Risk Levels (see MRL) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9, 16, 24, 275 
 

miscarriages . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13, 19, 103, 160, 179
 

monkey . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18, 90, 106, 110, 111, 116, 180
 

mosquito . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 213 
 

moss . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 227 
 

MRL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16-20, 24, 84, 105, 110, 112, 161, 176, 177, 179, 181, 269
 

MRLs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9, 16, 24, 25, 272, 275
 

musculoskeletal effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 80, 89, 109 
 

National Human Adipose Tissue Survey . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 202, 236, 241
 

National Priorities List (see NPL) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1, 199 
 

Native American . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 188, 231, 242, 244
 

neoplastic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 114, 271
 

neurobehavioral . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102, 157, 174, 176, 181, 186
 

neurodevelopmental . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 110, 180
 

NIOSH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9, 190, 238, 253, 271, 275
 

NOAEL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20, 23, 26, 73, 77, 78, 80, 83, 85, 100, 102, 108, 112
 

NOAELs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23, 24 
 

NOES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 238 
 

non-Hodgkin’s lymphoma . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 114 
 

no-observed-adverse-effect level . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20 
 

NPL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1, 199, 200, 202, 206, 210, 212, 242, 249, 250
 

octachlorostyrene . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25, 196 
 

ocular effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25, 93 
 

ovarian . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14, 18, 19, 103-106, 122, 157, 176, 179, 180
 

ovaries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6, 14, 18, 104-106, 133, 157, 161, 232, 269
 

ovary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12, 18, 19, 104, 151
 

parathyroid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6, 15, 81, 113, 116, 157, 161, 176, 178
 

paresthesia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16, 100, 181
 

particulate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201, 211, 216, 219, 220, 222, 246, 257
 

particulates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201, 210, 211
 

partition coefficients .............................................................................. 149, 191
 

PBPD . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145 
 

PBPK . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 144-147, 184, 185, 187
 

pembe yara . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6, 13, 16, 75, 91, 108, 159, 164, 169, 183
 

pentachlorobenzene . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 119, 123, 124, 137, 138, 141, 154, 156, 163, 182, 217-219
 

pentachlorophenol . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11, 89, 137-139, 141, 189, 193, 203-205, 208, 209, 265-267
 

pentachlorothiophenol ............................................................. 123, 124, 137-139, 153, 154
 

pharmacodynamic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145 
 

pharmacokinetic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 130, 143-147, 150, 176-180
 

photolysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 217 
 

phototoxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14, 92 
 

physiologically based pharmacodynamic (see PBPD) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145 
 

physiologically based pharmacokinetic (see PBPK) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 144, 145, 147
 

pig . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 229 
 

porphyria . . . . . . . . . . . . . . . . . . . . 4, 13-15, 25, 74, 81-84, 87, 91-93, 96, 103, 108, 109, 113, 126, 148, 153, 154, 156, 164-167,
 

169, 172-174, 176-178, 183, 256
 

precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201, 220
 

progesterone . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14, 18, 19, 88, 90, 104-107, 157, 165
 

public health . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1, 4, 8, 9, 11, 23, 24, 157, 173, 174, 187, 209, 244, 253, 264, 274
 

RCRA ................................................................................. 246, 270, 273, 275
 

reductive dechlorination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123, 152
 

reference dose (see RfD) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 269, 275
 
HEXACHLOROBENZENE D-4 

APPENDIX D 

regulations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8, 9, 196, 246, 269-272, 275
 

renal effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12, 26, 84, 86 
 

reportable quantity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 273 
 

reproductive effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14, 18, 73, 88, 103, 108, 109, 117, 179, 269
 

Resource Conservation and Recovery Act (see RCRA) ............................................... 246, 270, 275
 

RfD . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 272, 275
 

salmon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 225, 233
 

sea . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 87, 131, 214, 226-228, 235
 

sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11, 201, 202, 208, 210-212, 214, 216, 217, 219, 223, 224, 228, 246, 249, 260
 

selenium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 162 
 

serum . . . . . . . . . . . . . . . . . . . . 18-20, 25, 26, 72, 78, 81, 85, 87-90, 93, 96-99, 104-107, 109, 111-113, 125, 126, 132-135, 149,
 

157, 162-164, 180, 182, 183, 218, 233-238, 253, 255
 

shrimp . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 224 
 

smelt . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 214, 215, 234, 247, 248
 

snail . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 213, 228
 

soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2, 3, 7, 9, 11, 201-203, 208-215, 217, 223, 224, 232, 241, 242, 246-250, 259, 260, 265
 

solubility . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11, 191, 210
 

sportfish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 233 
 

Superfund . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 196, 225, 226
 

surface water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3, 201, 202, 211, 216, 217, 222
 

T3 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 87-90 
 

T4 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26, 87-90, 102 
 

TCHQ . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123, 153, 154, 166
 

tetrachlorobenzene . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123, 137-139, 217-219
 

tetrachlorocatechol . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 139 
 

tetrachlorohydroquinone ....................................................................... 137-139, 141
 

thyroid . . . . . . . . . . . . . . . . . . . . . . 5, 6, 12, 13, 15, 16, 26, 74, 77, 87-89, 112, 113, 115, 133, 151, 157, 161, 166, 167, 174, 176­

178, 188
 

thyroid stimulating hormone (see TSH) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26 
 

thyroxine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12, 26, 89, 90, 157 
 

time-weighted average (see TWA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 275 
 

total diet studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 233, 241
 

toxicokinetic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23, 185 
 

Toxics Release Inventory (see TRI) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 194, 202
 

transport . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 157, 172, 199, 203, 210, 212, 246, 247
 

tremors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17, 76, 100, 101, 148, 159, 176, 177, 181
 

TRI . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 194, 202, 206, 207, 210, 217, 246
 

trichlorophenol .................................................................................. 123, 138
 

triiodothyronine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 90 
 

TSH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26, 72, 88 
 

tumor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 74, 97, 110, 113-116, 155, 156, 168, 178, 187
 

tumors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 16, 74, 75, 94, 107, 112-116, 155, 168, 178
 

Turkey . . . . . . . . . . . . . . . . . . . . . . 4, 12, 15, 16, 75, 81, 82, 87, 91, 100, 103, 108, 113, 119, 126, 159, 169, 172, 177, 182, 241,
 

243
 

TWA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 271, 275
 

USDA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 272, 275
 

vapor phase . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 199, 210, 215, 216, 246
 

vapor pressure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 119, 191, 210, 220
 

volatilization ............................................................. 201, 203, 205, 211-213, 217, 241, 246
 

water . . . . . . . . . . . . . . . . . . . . 2, 3, 5, 7, 9, 11, 82, 87, 102, 118, 128, 131, 158, 171, 189, 191, 196, 197, 199, 201, 202, 206-208,
 

210, 211, 213, 214, 216-218, 221-223, 225, 233, 242, 246, 247, 249, 250, 257-260, 262, 263, 265,
 

269-274