1,3-butadiene: human health aspects · 1,3-butadiene: human health aspects please note that the...

This report contains the collective views of an international group of experts and does notnecessarily represent the decisions or the stated policy of the United Nations EnvironmentProgramme, the International Labour Organization, or the World Health Organization.

Concise International Chemical Assessment Document 30

1,3-BUTADIENE: HUMAN HEALTH ASPECTS

Please note that the layout and pagination of this pdf file are not identical to those of theprinted CICAD

First draft prepared by K. Hughes, M.E. Meek, M. Walker, and R. Beauchamp,Health Canada, Ottawa, Canada

Published under the joint sponsorship of the United Nations Environment Programme, theInternational Labour Organization, and the World Health Organization, and produced within theframework of the Inter-Organization Programme for the Sound Management of Chemicals.

World Health OrganizationGeneva, 2001

The International Programme on Chemical Safety (IPCS), established in 1980, is a joint ventureof the United Nations Environment Programme (UNEP), the International Labour Organization (ILO),and the World Health Organization (WHO). The overall objectives of the IPCS are to establish thescientific basis for assessment of the risk to human health and the environment from exposure tochemicals, through international peer review processes, as a prerequisite for the promotion of chemicalsafety, and to provide technical assistance in strengthening national capacities for the sound managementof chemicals.

The Inter-Organization Programme for the Sound Management of Chemicals (IOMC) wasestablished in 1995 by UNEP, ILO, the Food and Agriculture Organization of the United Nations, WHO,the United Nations Industrial Development Organization, the United Nations Institute for Training andResearch, and the Organisation for Economic Co-operation and Development (ParticipatingOrganizations), following recommendations made by the 1992 UN Conference on Environment andDevelopment to strengthen cooperation and increase coordination in the field of chemical safety. Thepurpose of the IOMC is to promote coordination of the policies and activities pursued by the ParticipatingOrganizations, jointly or separately, to achieve the sound management of chemicals in relation to humanhealth and the environment.

WHO Library Cataloguing-in-Publication Data

1,3-Butadiene : human health aspects.

(Concise international chemical assessment document ; 30)

1.Butadienes - toxicity 2.Risk assessment 3.Environmental exposure4.Occupational exposure I.International Programme on Chemical SafetyII.Series

ISBN 92 4 153030 8 (NLM Classification: QD 305.H7) ISSN 1020-6167

The World Health Organization welcomes requests for permission to reproduce or translate itspublications, in part or in full. Applications and enquiries should be addressed to the Office of Publications,World Health Organization, Geneva, Switzerland, which will be glad to provide the latest information onany changes made to the text, plans for new editions, and reprints and translations already available.

©World Health Organization 2001

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The designations employed and the presentation of the material in this publication do not imply theexpression of any opinion whatsoever on the part of the Secretariat of the World Health Organizationconcerning the legal status of any country, territory, city, or area or of its authorities, or concerning thedelimitation of its frontiers or boundaries.

The mention of specific companies or of certain manufacturers’ products does not imply that they areendorsed or recommended by the World Health Organization in preference to others of a similar naturethat are not mentioned. Errors and omissions excepted, the names of proprietary products aredistinguished by initial capital letters.

The Federal Ministry for the Environment, Nature Conservation and Nuclear Safety, Germany,provided financial support for the printing of this publication.

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TABLE OF CONTENTS

FOREWORD . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

1. EXECUTIVE SUMMARY . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

2. IDENTITY AND PHYSICAL/CHEMICAL PROPERTIES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

3. ANALYTICAL METHODS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

4. SOURCES OF HUMAN EXPOSURE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6

4.1 Natural sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 64.2 Anthropogenic sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6

4.3 Production and uses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7

5. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION . . . . . . . . . . . . . . . . . . . . . . 7

5.1 Air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 75.2 Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 85.3 Sediment and soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

5.4 Biota . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 85.5 Environmental modelling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

6. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9

6.1 Environmental levels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 96.1.1 Ambient air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 96.1.2 Surface water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 96.1.3 Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9

6.2 Human exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 96.2.1 Indoor air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 96.2.2 Drinking-water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106.2.3 Food . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106.2.4 Consumer products . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106.2.5 Occupational exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10

7. COMPARATIVE KINETICS AND METABOLISM IN LABORATORY ANIMALS ANDHUMANS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11

8. EFFECTS ON LABORATORY MAMMALS AND IN VITRO TEST SYSTEMS . . . . . . . . . . . . . . . . . . . . . . . . . . 13

8.1 Single exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 138.2 Irritation and sensitization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 138.3 Repeated exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 138.4 Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 138.5 Genotoxicity and related end-points . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 188.6 Reproductive toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23

8.6.1 Effects on fertility . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 238.6.2 Developmental toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24

8.7 Immunotoxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24


iv

9. EFFECTS ON HUMANS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24

9.1 Clinical studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 249.2 Epidemiological studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25

9.2.1 Cancer . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 259.2.2 Non-neoplastic effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 289.2.3 Genotoxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28

10. EVALUATION OF HEALTH EFFECTS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29

10.1 Hazard identification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2910.1.1 Carcinogenicity and genotoxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2910.1.2 Non-neoplastic effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32

10.2 Exposure–response assessment and criteria for setting tolerable concentrations orguidance values . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3310.2.1 Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 33

10.2.1.1 Epidemiological data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3310.2.1.2 Data from studies in experimental animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34

10.2.2 Non-neoplastic effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3510.3 Sample exposure and risk characterization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36

10.3.1 Sample exposure characterization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36

10.3.2 Sample risk characterization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3610.4 Uncertainties and degree of confidence in human health hazard characterization and

sample risk characterization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37

11. PREVIOUS EVALUATIONS BY INTERNATIONAL BODIES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40

REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41

APPENDIX 1 — SOURCE DOCUMENT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 49

APPENDIX 2 — CICAD PEER REVIEW . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50

APPENDIX 3 — CICAD FINAL REVIEW BOARD . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 51

APPENDIX 4 — QUANTITATION OF EXPOSURE–RESPONSE FOR CRITICAL EFFECTS ASSOCIATEDWITH EXPOSURE TO 1,3-BUTADIENE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52

INTERNATIONAL CHEMICAL SAFETY CARD . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 68

RÉSUMÉ D’ORIENTATION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 70

RESUMEN DE ORIENTACIÓN . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 72

1,3-Butadiene: Human health aspects

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FOREWORD

Concise International Chemical AssessmentDocuments (CICADs) are the latest in a family ofpublications from the International Programme onChemical Safety (IPCS) — a cooperative programme ofthe World Health Organization (WHO), the InternationalLabour Organization (ILO), and the United NationsEnvironment Programme (UNEP). CICADs join theEnvironmental Health Criteria documents (EHCs) asauthoritative documents on the risk assessment ofchemicals.

International Chemical Safety Cards on therelevant chemical(s) are attached at the end of theCICAD, to provide the reader with concise informationon the protection of human health and on emergencyaction. They are produced in a separate peer-reviewedprocedure at IPCS. They may be complemented byinformation from IPCS Poison Information Monographs(PIM), similarly produced separately from the CICADprocess.

CICADs are concise documents that provide sum-maries of the relevant scientific information concerningthe potential effects of chemicals upon human healthand/or the environment. They are based on selectednational or regional evaluation documents or on existingEHCs. Before acceptance for publication as CICADs byIPCS, these documents undergo extensive peer reviewby internationally selected experts to ensure theircompleteness, accuracy in the way in which the originaldata are represented, and the validity of the conclusionsdrawn.

The primary objective of CICADs is characteri-zation of hazard and dose–response from exposure to achemical. CICADs are not a summary of all available dataon a particular chemical; rather, they include only thatinformation considered critical for characterization of therisk posed by the chemical. The critical studies are,however, presented in sufficient detail to support theconclusions drawn. For additional information, thereader should consult the identified source documentsupon which the CICAD has been based.

Risks to human health and the environment willvary considerably depending upon the type and extentof exposure. Responsible authorities are stronglyencouraged to characterize risk on the basis of locallymeasured or predicted exposure scenarios. To assist thereader, examples of exposure estimation and riskcharacterization are provided in CICADs, wheneverpossible. These examples cannot be considered asrepresenting all possible exposure situations, but areprovided as guidance only. The reader is referred to EHC

1701 for advice on the derivation of health-basedtolerable intakes and guidance values.

While every effort is made to ensure that CICADsrepresent the current status of knowledge, new informa-tion is being developed constantly. Unless otherwisestated, CICADs are based on a search of the scientificliterature to the date shown in the executive summary. Inthe event that a reader becomes aware of new informa-tion that would change the conclusions drawn in aCICAD, the reader is requested to contact IPCS to informit of the new information.

Procedures

The flow chart shows the procedures followed toproduce a CICAD. These procedures are designed totake advantage of the expertise that exists around theworld — expertise that is required to produce the high-quality evaluations of toxicological, exposure, and otherdata that are necessary for assessing risks to humanhealth and/or the environment. The IPCS Risk Assess-ment Steering Group advises the Co-ordinator, IPCS, onthe selection of chemicals for an IPCS risk assessment,whether a CICAD or an EHC is produced, and whichinstitution bears the responsibility of the documentproduction, as well as on the type and extent of theinternational peer review.

The first draft is based on an existing national,regional, or international review. Authors of the firstdraft are usually, but not necessarily, from the institutionthat developed the original review. A standard outlinehas been developed to encourage consistency in form.The first draft undergoes primary review by IPCS andone or more experienced authors of criteria documents inorder to ensure that it meets the specified criteria forCICADs.

The draft is then sent to an international peerreview by scientists known for their particular expertiseand by scientists selected from an international rostercompiled by IPCS through recommendations from IPCSnational Contact Points and from IPCS ParticipatingInstitutions. Adequate time is allowed for the selectedexperts to undertake a thorough review. Authors arerequired to take reviewers’ comments into account andrevise their draft, if necessary. The resulting second draft

1 International Programme on Chemical Safety (1994)Assessing human health risks of chemicals: derivationof guidance values for health-based exposure limits.Geneva, World Health Organization (EnvironmentalHealth Criteria 170).


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S E L E C T I O N O F H I G H Q U A L I T YN A T I O N A L / R E G I O N A L

A S S E S S M E N T D O C U M E N T ( S )

CICAD PREPARATION FLOW CHART

F I R S T D R A F T

P R E P A R E D

REVIEW BY IPCS CONTACT POINTS/SPECIALIZED EXPERTS

FINAL REVIEW BOARD 2

FINAL DRAFT 3

EDITING

APPROVAL BY DIRECTOR, IPCS

PUBLICATION

SELECTION OF PRIORITY CHEMICAL

1 Taking into account the comments from reviewers.2 The second draft of documents is submitted to the Final Review Board together with the reviewers’ comments.3 Includes any revisions requested by the Final Review Board.

REVIEW OF COMMENTS (PRODUCER/RESPONSIBLE OFFICER),PREPARATION

OF SECOND DRAFT 1

P R I M A R Y R E V I E W B Y I P C S

( REVISIONS AS NECESSARY)


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is submitted to a Final Review Board together with thereviewers’ comments.

A consultative group may be necessary to adviseon specific issues in the risk assessment document.

The CICAD Final Review Board has severalimportant functions:

– to ensure that each CICAD has been subjected toan appropriate and thorough peer review;

– to verify that the peer reviewers’ comments have

been addressed appropriately;– to provide guidance to those responsible for the

preparation of CICADs on how to resolve anyremaining issues if, in the opinion of the Board, theauthor has not adequately addressed all commentsof the reviewers; and

– to approve CICADs as international assessments.

Board members serve in their personal capacity, not asrepresentatives of any organization, government, orindustry. They are selected because of their expertise inhuman and environmental toxicology or because of theirexperience in the regulation of chemicals. Boards arechosen according to the range of expertise required for ameeting and the need for balanced geographicrepresentation.

Board members, authors, reviewers, consultants,and advisers who participate in the preparation of aCICAD are required to declare any real or potentialconflict of interest in relation to the subjects underdiscussion at any stage of the process. Representativesof nongovernmental organizations may be invited toobserve the proceedings of the Final Review Board.Observers may participate in Board discussions only atthe invitation of the Chairperson, and they may notparticipate in the final decision-making process.


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1. EXECUTIVE SUMMARY

This CICAD on 1,3-butadiene was prepared by theEnvironmental Health Directorate of Health Canadabased on documentation prepared concurrently as partof the Priority Substances Program under the CanadianEnvironmental Protection Act (CEPA). The objectiveof health assessments on Priority Substances underCEPA is to assess the potential effects of indirectexposure in the general environment on human health.Data identified as of the end of April 1998 wereconsidered in this review. Information on the nature ofthe peer review and availability of the source documentis presented in Appendix 1. Information on the peerreview of this CICAD is presented in Appendix 2. ThisCICAD was approved as an international assessment ata meeting of the Final Review Board, held in Helsinki,Finland, on 26–29 June 2000. Participants at the FinalReview Board meeting are listed in Appendix 3. TheInternational Chemical Safety Card (ICSC 0017) for 1,3-butadiene, produced by the International Programme onChemical Safety (IPCS, 1993), has also been reproducedin this document.

1,3-Butadiene (CAS No. 106-99-0) is a product ofincomplete combustion resulting from natural processesand human activity. It is also an industrial chemical usedprimarily in the production of polymers, including poly-butadiene, styrene-butadiene rubbers and lattices, andnitrile-butadiene rubbers. 1,3-Butadiene enters theenvironment from exhaust emissions from gasoline- anddiesel-powered vehicles, from non-transportation fuelcombustion, from biomass combustion, and from indus-trial on-site uses.

While 1,3-butadiene is not persistent, it is ubiqui-tous in the urban environment because of its widespreadcombustion sources. The highest atmosphericconcentrations have been measured in air in cities andclose to industrial sources.

The general population is exposed to 1,3-butadieneprimarily through ambient and indoor air. In comparison,other media, including food and drinking-water, contrib-ute negligibly to exposure to 1,3-butadiene. Tobaccosmoke may contribute significant amounts of 1,3-buta-diene.

Metabolism of 1,3-butadiene appears to be qualita-tively similar across species, although there are quantita-tive differences in the amounts of putatively toxic metab-olites formed; mice appear to oxidize 1,3-butadiene to themonoepoxide, and subsequently the diepoxide,metabolite to a greater extent than do rats or humans.However, there may also be interindividual variation in

metabolic capability for 1,3-butadiene in humans, relatedto genetic polymorphism for relevant enzymes.

1,3-Butadiene is of low acute toxicity in experimen-tal animals. However, long-term exposure to 1,3-butadiene was associated with the development ofovarian atrophy at all concentrations tested in mice.Other effects in the ovaries have also been observed inshorter-term studies. Atrophy of the testes was alsoobserved in male mice at concentrations greater thanthose associated with effects in females. Based onlimited available data, there is no conclusive evidencethat 1,3-butadiene is teratogenic in experimental animalsfollowing maternal or paternal exposure or that it inducessignificant fetal toxicity at concentrations below thosethat are maternally toxic.

1,3-Butadiene also induced a variety of effects onthe blood and bone marrow of mice; although data arelimited, similar effects have not been observed in rats.

Inhaled 1,3-butadiene is a potent carcinogen inmice, inducing tumours at multiple sites at all concentra-tions tested in all identified studies. 1,3-Butadiene wasalso carcinogenic in rats at all exposure levels in the onlyrelevant study available; although only much higherconcentrations were tested in rats than in mice, ratsappear to be the less sensitive species, based oncomparison of tumour incidence data. The greatersensitivity in mice than in rats to induction of theseeffects by 1,3-butadiene is likely related to speciesdifferences in metabolism to the active epoxidemetabolites.

1,3-Butadiene is mutagenic in somatic cells of bothmice and rats, although the mutagenic potency wasgreater in mice than in rats. Similarly, 1,3-butadieneinduced other genetic damage in somatic cells of mice,but not in those of rats. 1,3-Butadiene was also consis-tently genotoxic in germ cells of mice, but not in thesingle assay in rats identified. However, there were noapparent differences in species sensitivity to geneticeffects induced by epoxide metabolites of 1,3-butadiene.There is also limited evidence from occupationallyexposed populations that 1,3-butadiene is genotoxic inhumans, inducing mutagenic and clastogenic damage insomatic cells.

An association between exposure to 1,3-butadienein the occupational environment and leukaemia fulfilsseveral of the traditional criteria for causality. In thelargest and most comprehensive study conducted todate, involving a cohort of workers from multiple plants,mortality due to leukaemia increased with estimatedcumulative exposure to 1,3-butadiene in the styrene-butadiene rubber industry; this association remained


5

after controlling for exposure to styrene and benzeneand was strongest in those subgroups with highestpotential exposure. Similarly, an association betweenexposure to 1,3-butadiene and leukaemia was observedin an independently conducted case–control study oflargely the same population of workers. However, therewas no increase in mortality due to leukaemia inbutadiene monomer production workers who were notconcomitantly exposed to some of the other substancespresent in the styrene-butadiene rubber industry,although there was some limited evidence of anassociation with mortality due to lymphosarcoma andreticulosarcoma in some subgroups.

The available epidemiological and toxicologicaldata provide evidence that 1,3-butadiene is carcinogenicin humans and may also be genotoxic in humans. Thecarcinogenic potency (the concentration associated witha 1% increase in mortality due to leukaemia) was deter-mined to be 1.7 mg/m3, based on the results of the largestwell conducted epidemiological investigation in exposedworkers. This value is similar to the lower end of therange of tumorigenic concentrations determined on thebasis of studies in rodents. 1,3-Butadiene also inducedreproductive toxicity in experimental animals. As ameasure of its potency to induce reproductive effects, abenchmark concentration of 0.57 mg/m3 was derived forovarian toxicity in mice.

Although the health effects associated with expo-sure to 1,3-butadiene and the mode of action for induc-tion of these effects have been extensively investigated,there continues to be considerable research on this sub-stance in an effort to address some of the uncertaintiesassociated with the database.

2. IDENTITY AND PHYSICAL/CHEMICALPROPERTIES

1,3-Butadiene (H2C=CH@CH=CH2) is also known asbutadiene, ",(-butadiene, buta-1,3-diene, bivinyl,divinyl, erythrene, vinylethylene, biethylene, andpyrrolylene. Its Chemical Abstracts Service (CAS)registry number is 106-99-0, and its Registry of ToxicEffects of Chemical Substances (RTECS) number isEI9275000.

At room temperature, butadiene is a colourless,flammable gas with a mild aromatic odour. The molecularweight of butadiene is 54.09 g/mol. It has a high vapourpressure (281 kPa at 25 °C), a vapour density of 1.9, amoderately low water solubility (735 mg/litre at 25 °C), alow boiling point (!4.4 °C), a low octanol/water partitioncoefficient (Kow 1.99) (Mackay et al., 1993), and a Henry’slaw constant of 7460 Pa@m3/mol (equivalent to anair/water partition coefficient, or dimensionless Henry’slaw constant, of 165.9).

Further chemical and physical characteristics ofbutadiene are given in the International Chemical SafetyCard reproduced in this document.

The conversion factor for butadiene in air is asfollows: 1 ppm = 2.21 mg/m3.

3. ANALYTICAL METHODS

Selected methods for the analysis of butadiene invarious matrices are listed in Table 1 (IARC, 1999). Gasdetection tubes can also be used to detect butadiene.

Table 1: Methods for analysis of butadiene (modified from IARC, 1999).

Sample matrix Sample preparationAssayprocedurea

Limit ofdetection Reference

Air Adsorb (charcoal); extract (carbon disulfide)Adsorb (coconut, charcoal); extract(dichloromethane)Adsorb on Perkin-Elmer ATD 400 packedwith polymeric or synthetic adsorbentmaterial; thermal desorption

GC/FIDGC/FID

GC/FID

200 µg/m3

0.2 mg/sample(5–25 litres)200 µg/m3

US OSHA, 1990NIOSH, 1994

UK HSE, 1992

Foods and plasticfood-packing material

Dissolve (dimethylacetamide) or melt; injectheadspace sample

GC/MS-SIM ~1 µg/kg Startin & Gilbert, 1984

Plastics, liquid foods Dissolve in o-dichlorobenzene; injectheadspace sample

GC/FID 2–20 µg/kg US FDA, 1987

Solid foods Cut or mash sample; inject headspacesample

GC/FID 2–20 µg/kg US FDA, 1987

a Abbreviations: GC/FID: gas chromatography/flame ionization detection; GC/MS-SIM: gas chromatography/mass spectrometry withsingle-ion monitoring.


6

4. SOURCES OF HUMAN EXPOSURE

Data on sources and emissions from Canada, thesource country of the national assessment on which thisCICAD is based, are presented here as an example.Sources and patterns of emissions in other countries areexpected to be similar, although quantitative values mayvary.

Estimates of emissions of butadiene are highlyvariable, depending on the method of estimation and thequality of the data upon which they are based. TotalCanadian emissions for 1994 were estimated to rangebetween 12 917 and 41 622 tonnes (Environment Canada,1998). Major uncertainties are associated with estimatesfor combustion sources, notably forest fires.

4.1 Natural sources

Butadiene is released from biomass combustion,especially forest fires. Total global emissions of buta-diene from biomass combustion were estimated to be 770000 tonnes per year (Ward & Hao, 1992). Releases fromforest fires in Canada were estimated to range between3607 and 26 966 tonnes, which constituted 49.3% (rangeof estimates is 28–65%) of the total annual emissions ofbutadiene in Canada (CPPI, 1997). Although Altshuller etal. (1971) suggested that butadiene can be released fromnatural gas losses and diffusion through soil frompetroleum deposits, no data were identified on thispossible source.

4.2 Anthropogenic sources

All internal combustion engines may producebutadiene as a result of incomplete combustion. Theamount generated and released depends primarily on thecomposition of fuel, the type of engine, the emissioncontrol used (i.e., presence and efficiency of catalyticconverter), the operating temperature, and the age andstate of repair of the vehicle.1 Cyclohexane, 1-hexene,1-pentene, and cyclohexene have been identified asprimary fuel precursors for butadiene (Schuetzle et al.,1994). As well, very low levels of butadiene itself may bepresent in gasoline and in liquefied petroleum gas.

Butadiene can also enter the environment from anystage in the production, storage, use, transport, or dis-posal of products with residual, free, or unreacted buta-diene. Data on Canadian industrial emissions have been

collected for industrial processes, plastic productsindustries, refined petroleum and coal productsindustries, and chemical and chemical productsindustries as part of the National Pollutant ReleaseInventory (NPRI) (Environment Canada, 1996a, 1997).Emissions other than those reported to the NPRI mayoccur, including from combustion of other fuels (e.g.,natural gas, oil, and wood space heating), prescribedforest burning, cigarettes, waste incineration, releasesfrom polymer products, releases from the use anddisposal of products containing butadiene, and spillage(Ligocki et al., 1994; Environment Canada, 1996b; OECD,1996).

The following amounts of butadiene wereestimated to have been released into the Canadianenvironment in 1994 from key transportation and relatedsources (Environment Canada, 1996a; CPPI, 1997):3376–7401 tonnes from on-road gasoline- and diesel-powered motor vehicles (with about 45–89% of thosereleases from gasoline engines and 11–55% from dieselengines); 150–258 tonnes from aircraft; 84–1689 tonnesfrom off-road motor vehicles; 84 tonnes from lawn-mowers; 40 tonnes from the marine sector; and 17 tonnesfrom the rail sector.

In addition, data from NPRI for 1994 (EnvironmentCanada, 1996a) listed a total of 270.4 tonnes releasedfrom the chemical and chemical products industries. Ofthis, 270.3 tonnes were released into air, 0.058 tonnesinto water (St. Clair River, Ontario), and 0.002 tonnesonto land. There were releases of 17.5 tonnes into airfrom the plastic products industries. A total of 22.3tonnes was released from the refined petroleum and coalproducts industries, of which 22.2 tonnes were releasedinto air. Off-site transfer of wastes (material sent for finaldisposal or treatment prior to final disposal) fromindustrial sites in Canada in 1994 was estimated toinclude a total of 131.3 tonnes of butadiene, with 128.7tonnes being sent to incineration, 2.1 tonnes to landfill,and 0.5 tonnes to municipal sewage treatment plants(Environment Canada, 1996a). Based on 1995 NPRI data(Environment Canada, 1997), the amount of butadieneestimated to have been released into the Canadianenvironment was 225.8 tonnes from industrial on-siteuses, with 0.058 tonnes released into water, 0.002 tonnesinto land, and 225.4 tonnes into air. Releases into airincluded air fugitive releases (172.8 tonnes), air stackreleases (36.3 tonnes), air storage releases (4.8 tonnes),air spill releases (1.1 tonnes), and other air releases (10.4tonnes).

Based on data in NPRI, it was estimated that thetotal release of butadiene from fuel distribution in 1994was 24 tonnes (Environment Canada, 1996a), althoughgasoline and diesel fuel contain little or no butadiene(US EPA, 1989).

1 Personal communication from L.A. Graham, River RoadEnvironmental Technology Centre, Environment Canada,Ottawa, Ontario, to Commercial Chemicals EvaluationBranch, Environment Canada, Hull, Quebec, 1996.


7

CPPI (1997) estimated that releases into theCanadian environment in 1994 were 1191 tonnes fromprescribed forest burning, 3706 tonnes from wood spaceheating, 11 tonnes from natural gas/oil space heating,and 1–9 tonnes from cigarettes.

4.3 Production and uses

Butadiene is produced during the combustion oforganic matter in both natural processes and humanactivities. In addition, it is produced commercially for usein the chemical polymer industry.

Butadiene is purified by extraction from a crudepetroleum butadiene stream. In 1994, there was oneCanadian commercial producer of butadiene (located inSarnia, Ontario), with a domestic production of103.7 kilotonnes. Importation into Canada from the USAwas 1.7 kilotonnes in 1994. The Canadian domestic useof butadiene in 1994 amounted to 105.4 kilotonnes(98.3 kilotonnes for total domestic demand and 7.1 forexport sales) (Camford Information Services, 1995). In theUSA, total production in 1993 was 1.4 billion kilograms.1

According to data summarized in IARC (1999), in 1996,production of butadiene in China (Taiwan), France,Germany, Japan, the Republic of Korea, and the USAwas 129, 344, 673, 1025, 601, and 1744 kilotonnes,respectively.

The largest end use of butadiene in Canada is theproduction of polybutadiene rubber (51.4 kilotonnes;52.3% of total Canadian consumption for 1994) (CamfordInformation Services, 1995). Other derivatives producedinclude styrene-butadiene lattices (31.0 kilotonnes;31.5% of total Canadian consumption for 1994), nitrile-butadiene rubbers (10.0 kilotonnes; 10.2% for 1994),acrylonitrile-butadiene-styrene terpolymer (3.4 kilo-tonnes; 3.5% for 1994), and specialty styrene-butadienerubbers (2.5 kilotonnes; 2.5% of total Canadianconsumption for 1994).

Butadiene has a long history of use, notablyrelated to production of polymers. Several industrial andcommercial products are manufactured with it or maycontain it as a component. Examples include tires, carsealants, plastic bottles and food wrap, epoxy resins,lubricating oils, hoses, drive belts, moulded rubbergoods, adhesives, paint, latex foams for carpet backingor underpad, shoe soles, moulded toys/householdgoods, medical devices, and chewing gum (CEH-SRIInternational, 1994; OECD, 1996).

5. ENVIRONMENTAL TRANSPORT,DISTRIBUTION, AND TRANSFORMATION

5.1 Air

Since butadiene is released primarily to air, its fatein that medium is of primary importance. Butadiene is notexpected to persist in air, since it oxidizes rapidly withseveral oxidant species. Destruction of atmosphericbutadiene by the gas-phase reaction with photochemi-cally produced hydroxyl radicals is expected to be thedominant photo-initiated pathway. Products that can beformed include formaldehyde, acrolein, and furan.Destruction by nitrate radicals is expected to be a sig-nificant nighttime process in urban areas. Acrolein,trans-4-nitroxy-2-butenal, and 1-nitroxy-3-buten-2-onehave been identified as products of this reaction. Reac-tion with ozone is also rapid but less important thanreaction with hydroxyl radicals. The products of thereaction of butadiene with ozone are acrolein, formalde-hyde, acetylene, ethylene, formic acid, formic anhydride,carbon monoxide, carbon dioxide, hydrogen gas, hydro-peroxyl radical, hydroxyl radical, and 3,4-epoxy-1-butene(Atkinson et al., 1990; Howard et al., 1991; McKone etal., 1993; US EPA, 1993).

Average atmospheric half-lives for photo-oxidationof butadiene, based on measured as well as calculateddata, range from 0.24 to 1.9 days (Darnell et al., 1976;Lyman et al., 1982; Atkinson et al., 1984; Becker et al.,1984; Klöpffer et al., 1988; Howard et al., 1991; Mackay etal., 1993). However, half-lives for butadiene in air canvary considerably under different conditions.Estimations for atmospheric residence time in several UScities ranged from 0.4 h under clear skies at night in thesummer to 2000 h (83 days) under cloudy skies at nightin the winter. Daytime residence times for different citieswithin a given season varied by factors of 2–3. Nighttimeresidence times varied by larger factors. The differencesbetween summer and winter conditions were large at allsites, with winter residence times 10–30 times greaterthan summer residence times (US EPA, 1993). Because ofthe long residence times under some conditions,especially in winter under cloudy conditions, there is apossibility of day-to-day carryover. Nonetheless, giventhe generally short daytime residence times, the netatmospheric lifetime of butadiene is short, and there isgenerally limited potential for long-range transport ofthis compound.

It is predicted from its physical/chemical propertiesthat when butadiene is released into air, almost all of itwill exist in the vapour phase in the atmosphere(Eisenreich et al., 1981; Environment Canada, 1998). Wetand dry deposition are not expected to be important as

1 Hazardous Substances Databank, National Library ofMedicine’s TOXNET system, searched 10 December1999.


8

transfer processes. Evaporation from rain may be rapid,and the compound is returned to the atmosphererelatively quickly unless it is leached into the soil.

5.2 Water

Volatilization, biodegradation, and oxidation bysinglet oxygen are the most prominent processesinvolved in determining the fate of butadiene in water.The estimated half-lives of butadiene by reaction inwater range from 4.2 to 28 days (Howard et al., 1991;Mackay et al., 1993).

5.3 Sediment and soil

The processes that are most prominent in deter-mining the environmental fate of butadiene in sedimentare biotic and abiotic degradation. The modelled half-lives of butadiene by reaction in sediment range from41.7 to 125 days (Mackay et al., 1993).

Based on its vapour pressure and its solubility,volatilization of butadiene from soil and other surfaces isexpected to be significant. Butadiene’s organic carbon/water partition coefficient indicates that it should notadsorb to soil particles to a great degree and would beconsidered moderately mobile (Kenaga, 1980; Swann etal., 1983). However, the rapid rate of volatilization andthe potential for degradation in soil suggest that it isunlikely that butadiene will leach into groundwater.Based on modelling predictions, the half-life of butadi-ene by reaction, given by Howard et al. (1991) andMackay et al. (1993), ranges from 7 to 41.7 days.

5.4 Biota

There are no measured bioconcentration factors.

Butadiene is metabolized by the mixed-function oxidasesystem in higher organisms, which contributes to theexpected lack of accumulation by many organisms.Estimated bioconcentration factors for butadiene in fishhave been reported to range from 4.6 to 19 (Lyman et al.,1982; OECD, 1996). Even though estimation methodslikely overestimate the true bioconcentration potentialfor a readily metabolized substance, they indicate thatbutadiene is not expected to bioconcentrate in aquaticorganisms or to biomagnify in the aquatic food chain.

There are no reported measurements of plant rootbioconcentration in soils. However, McKone et al. (1993)estimated the uptake of butadiene by roots from soilsolution to be 1.84 litres/kg, which is the ratio ofbutadiene concentration in root (mg/kg, fresh mass) tothe concentration in soil solution (mg/litre). The partitioncoefficient of butadiene concentration in roots (mg/kg,

fresh mass) to concentration in soil solids (mg/kg) wasestimated to range from 0.32 to 15 (dimensionless).

The partition coefficient of butadiene concentra-tion in whole plants (mg/kg, fresh mass) to its concentra-tion in soil solids (mg/kg) was estimated to range from0.1 to 2.9 (dimensionless). The steady-state plant/airpartition coefficient for foliar uptake of butadiene inplant leaves was estimated to be 0.63 m3/kg. There are noreported bioaccumulation data for any terrestrialinvertebrates.

5.5 Environmental modelling

Fugacity modelling was conducted to provide anoverview of key reaction, intercompartment, and advec-tion (movement out of a system) pathways for butadieneand of its overall distribution in the environment. Asteady-state, non-equilibrium model (Level III fugacitymodelling) was run using the methods developed byMackay (1991) and Mackay and Paterson (1991).Assumptions, input parameters, and results arepresented in Environment Canada (1998). Based onbutadiene’s physical/chemical properties, Level IIIfugacity modelling predicts that: • when butadiene is released into air, the distribution

of mass is almost 100% in air, with very smallamounts in soil and water;

• when butadiene is released into water, the distribu-tion of mass is 99.0% in water, with small amountsin air;

• when butadiene is released into soil, the distribu-tion of mass is 38.6% in soil, 59.3% in air, and 2.1%in water.

Modelling predictions do not purport to reflect

actual expected measurements in the environment butrather indicate the broad characteristics of the fate of thesubstance in the environment and its generaldistribution between media. Thus, when butadiene isdischarged into air or water, most of it is expected to befound in the medium receiving the discharge directly. Forexample, if butadiene is discharged into air, almost all ofit will exist in the atmosphere, where it will react rapidlyand will also be transported away. If butadiene isdischarged to water, it will react in water, and some willalso evaporate into air. If butadiene is discharged to soil,most will be present in air or soil, where it will react(Mackay et al., 1993; Environment Canada, 1998).


9

6. ENVIRONMENTAL LEVELS ANDHUMAN EXPOSURE

Data on environmental levels and human exposurefrom Canada, the source country of the national assess-ment on which this CICAD is based, are presented hereas a basis for the sample risk characterization. Patterns ofexposure in other countries are expected to be similar,although quantitative values may vary.

6.1 Environmental levels

6.1.1 Ambient air

Butadiene was detected (detection limit 0.05 µg/m3)in 7314 (or 80%) of 9168 24-h samples collected between1989 and 1996 from 47 sites across Canada.1 The meanconcentration in all samples was 0.3 µg/m3 (in thecalculation of the mean, a value of one-half the detectionlimit was assumed for samples in which levels werebelow the detection limit), and the maximumconcentration measured was 14.1 µg/m3. Concentrationsof butadiene in ambient air corresponding to the 50thand 95th percentiles were 0.21 and 1.0 µg/m3,respectively. Concentrations were generally higher inurban areas, with a mean exposure to 0.4 mg/m3 (95thpercentile 1.3 mg/m3) estimated as a “reasonable worst-case scenario,” based on data from four sites. Similarlevels were measured in smaller surveys in Canada (Bellet al., 1991, 1993; Hamilton-Wentworth, 1997; ConorPacific Environmental, 1998).2 In areas influenced byindustrial point sources of butadiene, concentrations inair were greater, with maximum and mean levels of 28 and0.62 mg/m3, respectively (95th percentile 6.4 mg/m3)being measured between 1 and 3 km from the source(MOEE, 1995).

Butadiene has also been detected in air in enclosedstructures. Concentrations of butadiene between 4 and49 µg/m3 were measured during the winter months of1994–1995 in Canadian underground parking garages(Environment Canada, 1994) because of its presence in

vehicle exhaust. Similarly, butadiene was frequentlydetected in samples from 10 parking structures in Cali-fornia, with the maximum concentration being 28 µg/m3

(Wilson et al., 1991). Butadiene has also been detected inurban road tunnels during rush hours in Australia (meanconcentration 28 µg/m3; Duffy & Nelson, 1996) andSweden (mean concentrations 17 µg/m3 and 25 µg/m3 intwo tunnels; Barrefors, 1996). Butadiene was measured atconcentrations ranging from 0.2 to 28 µg/m3 in 96 of 97 5-min air samples collected from a pumping island atrandomly identified self-service filling stations inCalifornia (Wilson et al., 1991). 6.1.2 Surface water

No data on concentrations of butadiene in Cana-dian lake, river, estuarine, or marine waters were identi-fied in the literature. Butadiene is being monitored ineffluents discharged into the St. Clair River from thebutadiene production plant in Sarnia, Ontario. It wasdetected only twice, at 2 and 5 µg/litre, in 2103 compositesamples of aqueous effluent taken every 4 h in 1996(detection limit 1 µg/litre). In daily sampling of effluentsfrom the four individual outfalls (detection limit 1 µg/litrein 736 samples and 50 µg/litre in 789 samples), butadienewas detected in only three samples, at concentrations of21, 80, and 130 µg/litre.3

6.1.3 Groundwater

Butadiene was detected but not quantified in agroundwater plume near a waste site in Quebec whererefinery oil residues and a variety of organic chemicalshad been dumped (Pakdel et al., 1992).

6.2 Human exposure

6.2.1 Indoor air

In available surveys in Canada, 1,3-butadiene wasdetected up to 6 times more frequently in indoor air inhomes than in corresponding samples of outdoor air,with concentrations being up to 10-fold higher indoorsthan outdoors (Bell et al., 1993; Hamilton-Wentworth,1997; Conor Pacific Environmental, 1998).4 Concentra-tions in air of indoor environments are highly variableand depend largely on individual activities and circum-

1 Unpublished data on butadiene levels in Canada fromNational Air Pollution Surveillance program, provided byT. Dann, River Road Environmental Technology Centre,Environment Canada, Ottawa, Ontario, to CommercialChemicals Evaluation Branch, Environment Canada, Hull,Quebec, April 1997.

2 Also letter dated 28 August 1996 from P. Steer, Scienceand Technology Branch, Ontario Ministry ofEnvironment and Energy, to J. Sealy, Health Canada, re.1,3-butadiene and chloroform data (File No.1E080149.MEM).

3 Personal communication from H. Michelin, Bayer Inc.,Sarnia, Ontario, to Commercial Chemicals EvaluationBranch, Environment Canada, Hull, Quebec, 1997.

4 Also personal communication dated 24 December 1997from X.-L. Cao to Health Canada, re. method detectionlimits for 24-h air samples from multimedia exposure pilotstudy (File No. MDL.XLS).


10

stances, including the use of consumer products (e.g.,cigarettes), the infiltration of vehicle exhaust from nearbytraffic and possibly from attached garages, and cookingactivities involving heated fats and oils (see section6.2.3). While data are inadequate to determine therelative contributions of each of these potential indoorsources, the highest concentrations of butadiene inindoor air in Canada have generally been detected inindoor environments contaminated with environmentaltobacco smoke (ETS). In a survey of 94 homes acrossCanada, the mean level in “non-smoking” homes was<1 µg/m3 (data censored by considering levels to be one-half the detection limit in samples in which butadienewas not detected), compared with a mean of 2.5 µg/m3

(data censored) in homes where smoking was present(Conor Pacific Environmental, 1998). Similarly, meanconcentrations in indoor air from “non-smoking” loca-tions in Windsor, Ontario, ranged from 0.3 to 1.6 µg/m3,while mean levels in “smoking” locations ranged from 1.3to 18.9 µg/m3. At non-residential indoor sampling sites inWindsor, the frequency of detection of butadiene was75–100% where ETS was present (Bell et al., 1993).

6.2.2 Drinking-water

There are no data available concerning the pres-ence of butadiene in drinking-water. In an investigationon whether the use of polybutylene pipe in water distri-bution systems is likely to result in the contamination ofdrinking-water with butadiene, Cooper1 did not detectthe substance in water from these types of pipes (nofurther information was presented in the secondaryaccount [CARB, 1992] of this study).

6.2.3 Food

There are no data available concerning thepresence or concentrations of butadiene in food inCanada. In the USA, the migration of butadiene fromrubber-modified plastic containers to food was inves-tigated by McNeal & Breder (1987). Butadiene wasdetected in some of the containers, but was generallynot detected in the foods (detection limits 1–5 ng/g).Similarly, in the United Kingdom, butadiene was notdetected (detection limit 0.2 ng/g) in five brands of softmargarine, although its presence was demonstrated (atconcentrations ranging from <5 to 310 ng/g) in theplastic containers (Startin & Gilbert, 1984). Butadiene hasbeen detected in the emissions from heated cooking oils,including Chinese rapeseed, peanut, soybean, andcanola oils, at levels ranging from 23 to 504 µg/m3

(Pellizzari et al., 1995; Shields et al., 1995).

6.2.4 Consumer products

Data on emissions of butadiene from potentialindoor sources such as styrene-butadiene rubber werenot identified.

Butadiene has been detected in both mainstreamsmoke and sidestream smoke from cigarettes in Canadaand the USA. For 18 brands of Canadian cigarettes, themean butadiene content ranged from 14.3 to 59.5 µg/cig-arette (overall mean concentration 30.0 µg/cigarette) inthe mainstream smoke and from 281 to 656 µg/cigarette(overall mean concentration 375 µg/cigarette) in the side-stream smoke, according to “preliminary” data (Labstat,Inc., 1995). The US DHHS (1989) reported that thevapour phase of mainstream smoke of non-filtercigarettes contained butadiene at levels of 25–40 µg/cig-arette. Brunnemann et al. (1989) measured butadienelevels ranging from 16 to 75 µg/cigarette in mainstreamsmoke from seven brands of cigarettes and levelsranging from 205 to 361 µg/cigarette in the sidestreamsmoke from six types of cigarettes. As discussed insection 6.2.1, the presence of ETS contributes toelevated levels of butadiene in indoor air.

6.2.5 Occupational exposure

Potential occupational exposure to butadiene canoccur in petroleum refining and related operations, pro-duction of butadiene monomer, production of butadiene-based polymers, or the manufacture of rubber and plas-tics products (IARC, 1999). Arithmetic mean concentra-tions in petroleum and petrochemical operations inseveral European countries ranged from 0.1 to 6.4 mg/m3

during 1984–1987 (IARC, 1999; European ChemicalsBureau, 2001). Based on occupational hygiene surveysof butadiene production facilities in the United Kingdom,personal airborne exposures are generally below a meanconcentration of 5 ppm (11 mg/m3), with most below 1ppm (2.2 mg/m3). In polymer manufacture in the UnitedKingdom, most time-weighted average exposures arebelow 2–3 ppm (4.4–6.6 mg/m3). Similar concentrationswere reported in other facilities in the European Union(IARC, 1999). In monomer production facilities in theUSA surveyed in 1985, arithmetic mean concentrationsranged from 1 to 277 mg/m3, while those in polymerproduction industries ranged from 0.04 to 32 mg/m3

(IARC, 1999).

1 Personal communication from R. Cooper, Department ofBiomedical and Environmental Health, School of PublicHealth, University of California, Berkeley, California,1989 (cited in CARB, 1992).


11

CH2 CH

CH2OH

CH

OH

RS

OH

1,3,4-trihydroxy-2-(N-acetylcysteinyl)butane

CH2 CH

OCH

CH2

O

1,2,3,4-diepoxybutane (DEB)

CH2 CH

CH2

OCH

1,2-epoxy-3-butene (EB)

CH2 CH

CH

CH2

OH OH1,3-dihydroxy-3-butene

CH2 CH

CH

2O

CH

OHOH1,2-dihydroxy-3,4-epoxybutane (EBdiol)

CH2 CH

CH

2O

CH


CH2 CH

CH

CH2

OH

RS

1-hydroxy-2-(N-acetylcysteinly)-3-butene (MII) via 1-hydroxy-2-(glutathionyl)-3-butene

CH2 CH

CH

CH2

RS

OH

RS CH2

CH2

CH

CH2

OH OH

1,2-dihydroxy-4-(N-acetylcysteinyl)butane (MI)

1-(N-acetylcysteinyl)-2-hydroxy-3-butene via 1-(glutathionyl)-2-hydroxy-3-butene

CH2 CH

CH

CH2

OH

OH

OH

OH

CH2 CH

CH

CH2

1,3-butadiene

CH2 CH

CH2

CO

H3-butenal

CH3 CH

CH

CH

O

crotonaldehyde

CH2 CH

CH

CH2

OH

OH OH

RS

1,3,4-trihydroxy-2-(N-acetylcysteinyl)butane erythritol

CH2 CH

CH

CH2

OH

OH

OH

OH

erythritol

CH2 CH

C CH3

O

acrolein

CH

C CH2

OHOH

RSOCH2

S-(1-hydroxy-3,4-epoxybut-2-yl)glutathione

CH

CH

CH

RSO

CH2

OHOHS-(2-hydroxy-3,4-epoxybut-1-yl)glutathione

CO2

+CO2

CO2

GSH

GT

GSH

GT

EH

P450

GSH

GT

GSH

GT

EH

EH

GSHGT

EH

P450

CH2 CH

CH2OH

CH

OH

RS

OH

1,3,4-trihydroxy-2-(N-acetylcysteinyl)butane

CH2 CH

OCH

CH2

O

1,2,3,4-diepoxybutane (DEB)

CH2 CH

CH2

OCH

1,2-epoxy-3-butene (EB)

CH2 CH

CH

CH2

OH OH1,3-dihydroxy-3-butene

CH2 CH

CH

2O

CH


CH2 CH

CH

2O

CH


CH2 CH

CH

CH2

OH

RS

1-hydroxy-2-(N-acetylcysteinly)-3-butene (MII) via 1-hydroxy-2-(glutathionyl)-3-butene

CH2 CH

CH

CH2

RS

OH

RS CH2

CH2

CH

CH2

OH OH

1,2-dihydroxy-4-(N-acetylcysteinyl)butane (MI)

1-(N-acetylcysteinyl)-2-hydroxy-3-butene via 1-(glutathionyl)-2-hydroxy-3-butene

CH2 CH

CH

CH2

OH

OH

OH

OH

CH2 CH

CH

CH2

1,3-butadiene

CH2 CH

CH2

CO

H3-butenal

CH3 CH

CH

CH

O

crotonaldehyde

CH2 CH

CH

CH2

OH

OH OH

RS

1,3,4-trihydroxy-2-(N-acetylcysteinyl)butane erythritol

CH2 CH

CH

CH2

OH

OH

OH

OH

erythritol

CH2 CH

C CH3

O

acrolein

CH

C CH2

OHOH

RSOCH2

S-(1-hydroxy-3,4-epoxybut-2-yl)glutathione

CH

CH

CH

RSO

CH2

OHOHS-(2-hydroxy-3,4-epoxybut-1-yl)glutathione

CO2

+CO2

CO2

GSH

GT

GSH

GT

EH

P450

GSH

GT

GSH

GT

EH

EH

GSHGT

EH

P450

Figure 1: Proposed metabolism of 1,3-butadiene.

7. COMPARATIVE KINETICS ANDMETABOLISM IN LABORATORY ANIMALS

AND HUMANS

The database on the toxicokinetics and metabolism

of butadiene is relatively extensive. The proposedmetabolism is outlined in Figure 1, based on the path-ways described by Henderson et al. (1993, 1996) andHimmelstein et al. (1997). Available data for the path-ways most extensively investigated indicate thatmetabolism is qualitatively similar among the variousspecies studied, although there may be quantitativedifferences in the amount of butadiene absorbed as wellas in metabolic rates and the proportion of metabolitesgenerated. These differences appear to be inconcordance with the observed variation in sensitivityto butadiene-induced toxic effects of the few strains ofrodent species tested to date, in that mice appear tometabolize a greater proportion of butadiene to activeepoxide metabolites than do rats. While less of thesemetabolites are also formed in samples of human tissues

in vitro than in those of mice, available data areinsufficient to characterize interindividual variability inhumans. Although there are known geneticpolymorphisms for a number of the enzymes involved inthe metabolism of butadiene, information on genotypewas not included in most investigations in humans.

Based on the metabolic pathways described inFigure 1, butadiene is first oxidized via cytochrome P-450enzymes (primarily P-450 2E1 in humans, although otherisoforms may also be involved, the relative contributionsof which vary between tissues and species) to themonoepoxide 1,2-epoxy-3-butene, or EB, which is subse-quently further oxidized via P-450 enzymes to thediepoxide 1,2,3,4-diepoxybutane, or DEB, or hydrolysedvia epoxide hydrolase (EH) to butenediol (1,2-dihydroxy-3-butene). The monoepoxide, the diepoxide, and thebutenediol may all be conjugated with glutathione (GSH)to form mercapturic acids, which are eventuallyeliminated in the urine. Hydrolysis of the diepoxide viaepoxide hydrolase or oxidation of the butenediol viacytochrome P-450 will result in the formation of the


12

monoepoxide diol (EBdiol). A small amount of butadienemay be converted to 3-butenal, which is subsequentlytransformed to crotonaldehyde (about 2–5% of theamount that is oxidized to the monoepoxide in humanliver microsomes [Duescher & Elfarra, 1994] ormicrosomes of kidney, lung, or liver of B6C3F1 mice[Sharer et al., 1992]). However, this pathway has notbeen extensively investigated, nor was crotonaldehydedetected in a sensitive analysis (using nuclear magneticresonance spectroscopy) of urinary metabolites of ratsand mice exposed to 13C-butadiene (Nauhaus et al.,1996).

Metabolism of butadiene and subsequent conver-sion of EB to DEB may also take place to a more limiteddegree in the bone marrow (e.g., Maniglier-Poulet et al.,1995) by means other than P-450 oxidation (possibly viamyeloperoxidase; Elfarra et al., 1996), based on in vitroobservations and the detection of the epoxides in thebone marrow of rodents (Thornton-Manning et al.,1995a, 1995b), although this potential pathway has notyet been extensively investigated. EB may also reactwith both myeloperoxidase and chloride to form achlorohydrin (1-chloro-2-hydroxy-3-butene) (Duescher &Elfarra, 1992). Metabolites arising from other possiblepathways have been identified in the urine of miceexposed to butadiene (including metabolites known tobe derived from metabolism of acrolein or acrylic acid)(Nauhaus et al., 1996), but no further research has yetbeen generated.

There is a substantial amount of evidence from invitro and in vivo investigations that B6C3F1 mice oxidizebutadiene to the monoepoxide via P-450 in the liver to agreater extent than do Sprague-Dawley rats and humans.Levels of EB in the blood and other tissues of mice weretwo- to eightfold higher than those in rats exposed tosimilar levels of butadiene (Bond et al., 1986; Himmel-stein et al., 1994, 1995; Bechtold et al., 1995; Thornton-Manning et al., 1997).

Available data also suggest that there are similarspecies differences in the amount of the diepoxideformed from oxidation of the monoepoxide. Levels ofDEB were 40- to 160-fold higher in blood and othertissues of B6C3F1 mice than in Sprague-Dawley ratsexposed to the same concentration of butadiene (Thorn-ton-Manning et al., 1995a, 1995b). While concentrationsof EB at various sites were similar in male and femalerats, levels of DEB were at least fivefold higher infemales than in males, which correlates with the greaterincidence of tumours in female rats. Although the mam-mary gland is a target tissue in rats, extended exposureto butadiene at 8000 ppm (17 696 mg/m3) for 10 days didnot result in any accumulation of DEB at this site(Thornton-Manning et al., 1998), which suggests that

DEB may not play a significant role in the induction ofmammary tumours in rats. Available in vitro data inhuman liver and lung samples suggest that humans alsoform less of the active metabolites of butadiene than domice (although somewhat varying results have beenreported with respect to the magnitude of the differencesbetween species) (Csanády et al., 1992; Duescher &Elfarra, 1994; Krause & Elfarra, 1997).

Although epoxide metabolites of butadiene areformed to a greater extent in mice than in rats or humans,they are also cleared via glutathione conjugation morerapidly in mice (Kreuzer et al., 1991; Sharer et al., 1992;Boogaard et al., 1996a, 1996b). Conversely, hydrolysis ofEB and DEB is greater in humans than in rats (based onin vitro data, as DEB has not been detected in tissues ofexposed humans), and hydrolysis of EB and DEB in ratsis in turn greater than that in mice (Csanády et al., 1992;Krause et al., 1997). In both humans and monkeys,removal of EB via hydrolysis appears to predominateover conjugation with glutathione, based on analysis ofurinary metabolites (Sabourin et al., 1992; Bechtold et al.,1994). Although hydrolysis of the epoxide metabolites isgenerally considered to be a detoxifying mechanism, itmay also lead to the formation of the diolepoxide, EBdiol,which is biologically reactive. However, no data wereidentified on species differences in the formation ofEBdiol via metabolism of both epoxide metabolites.

The formation of stable adducts of both the mono-epoxide and monoepoxide diol metabolites of butadienewith the N-terminal valine of haemoglobin has beenobserved in experimental animals and humans exposedto butadiene (Albrecht et al., 1993; Osterman-Golkar etal., 1993, 1996; Neumann et al., 1995; Sorsa et al., 1996b;Tretyakova et al., 1996; Pérez et al., 1997).1 Consistentwith the greater formation of epoxide metabolites, greaterconcentrations of haemoglobin–EB adducts weremeasured in mice than in rats exposed to the sameconcentration of butadiene. However, levels ofhaemoglobin–EB adducts in butadiene-exposed workers,although significantly elevated compared with levels innon-exposed workers, were considerably less thanwould be expected on the basis of results of studies inmice and rats (Osterman-Golkar et al., 1993). Based onobservations in rats and humans exposed to butadiene,levels of haemoglobin–EBdiol adducts are substantiallygreater than levels of haemoglobin–EB adducts(although it is noted that the same adduct can resultfrom binding with DEB). Metabolites of butadiene mayalso form adducts with DNA (see sections 8.5 and 9.2.3).

1 Also personal communication (correspondence dated25 March 1998) from J.A. Swenberg, University of NorthCarolina, Chapel Hill, NC, to Health Canada.


13

In addition to quantitative interspecies differencesin the metabolism of butadiene, there is also evidencethat there is significant variation within the human popu-lation. Indeed, although available data are inadequate toassess interindividual variation in metabolism, which hasbeen observed in in vitro investigations in microsomesfrom a small number of subjects (Boogaard & Bond,1996; Krause et al., 1997), there has been significantinterindividual variability in the extent of formation ofhaemoglobin adducts with butadiene metabolites inhuman populations (Neumann et al., 1995; Osterman-Golkar et al., 1996). Such variability is not unexpected, inview of the complexity of the metabolic pathwaysinvolved in the biotransformation of butadiene: i.e., thethree principal enzymatic processes that determine theextent of exposure to the putatively toxic epoxidemetabolites, namely formation via cytochrome P-450 2E1and removal via epoxide hydrolase and glutathioneconjugation. For example, the inducibility of cytochromeP-450 2E1 by low molecular weight compounds such asethanol is likely to contribute to interindividual variabil-ity in sensitivity. Moreover, genetic polymorphisms forglutathione-S-transferases and epoxide hydrolase mightalso contribute to considerable variation in sensitivity.While the influence of genotype for epoxide hydrolasehas not been well investigated (although data indicatethat hydrolysis of EB predominates over oxidation andglutathione conjugation in humans), interindividualsensitivity to the genetic effects of the epoxide metabo-lites in in vitro studies has been clearly related to geno-type for the glutathione-S-transferases (see section9.2.3).

8. EFFECTS ON LABORATORYMAMMALS AND IN VITRO TEST SYSTEMS

8.1 Single exposure

Although few data are available, butadieneappears to be of low acute toxicity in experimentalanimals, with reported LC50 values for rats and mice of>100 000 ppm (>221 000 mg/m3). Lowest LC50 values forbutadiene are reported for mice, at 117 000 ppm (256 000mg/m3) (duration not specified) (Batinka, 1966) and121 000 ppm (268 000 mg/m3) (2 h) (Shugaev, 1969). Thenervous system and the blood appear to be the principaltargets; however, in only one study were data sufficientto determine a lowest-observed-effect level (LOEL) of200 ppm (442 mg/m3) for haematological effects (Leavenset al., 1997). Exposure to butadiene for 7 h caused aconcentration-dependent depletion (by as much as 80%)of cellular non-protein sulfhydryl content of liver, lung,or heart in mice, with a LOEL of 100 ppm (221 mg/m3)

(Deutschmann & Laib, 1989). Depletion of non-proteinsulfhydryl content may inhibit detoxification of epoxidemetabolites vie glutathione conjugation.

8.2 Irritation and sensitization

No investigations in experimental animals on thepotential for irritation or sensitization of butadiene havebeen identified.

8.3 Repeated exposure

The majority of short-term and subchronic studieswere designed as either range-finding studiespreliminary to chronic bioassays or investigations ofpotential mechanisms of action for butadiene-inducedcancer and are not adequate for determination of criticaleffect levels. Effects on body weight were observed inB6C3F1 mice exposed to 625 ppm (1383 mg/m3) butadieneor more for 2 weeks; no histopathological changes werenoted at any concentration at or below 8000 ppm (17 696mg/m3) (NTP, 1984).

Haematological effects consistent with megalo-blastic anaemia and effects on bone marrow, includingalterations in stem cell development, have beenobserved in two strains of mice (B6C3F1 and NIH Swiss)exposed to 1000 or 1250 ppm (2212 or 2765 mg/m3)butadiene for up to 31 weeks (Irons et al., 1986a, 1986b;Leiderman et al., 1986; Bevan et al., 1996). Other effects,including decreased survival and body weight gain (withmales being more sensitive than females), altered organweights, and ovarian or testicular atrophy, have alsobeen observed in B6C3F1 mice exposed subchronicallyto similar or higher levels of butadiene (NTP, 1984;Bevan et al., 1996). In addition, an increased incidence ofa variety of tumours has been observed in B6C3F1 miceexposed to 625 ppm (1383 mg/m3) butadiene for as littleas 13 weeks (NTP, 1993) (see section 8.4). Althoughhistopathological changes and haematological effectswere reported in early studies in rats exposed to lowconcentrations (3 or 10 mg/m3) (Batinka, 1966; Ripp,1967; Nikiforova et al., 1969), these results were notconfirmed in more recent investigations of rats exposedfor up to 13 weeks to much higher concentrations (e.g.,17 600 mg/m3) (e.g., Crouch et al., 1979; Bevan et al.,1996). In view of the limitations of the studies in rats, it isnot possible to draw any conclusions regarding speciesdifferences in response to subchronic exposure tobutadiene.

8.4 Carcinogenicity

The carcinogenic potential of inhaled butadienehas been studied in two strains of mice and one strainof rats. Butadiene was a multi-site carcinogen in all


14

identified long-term experiments, inducing common andrare tumours in mice and rats, although there appear tobe marked species and strain differences in sensitivity.In an early bioassay by the National Toxicology Program(NTP, 1984) in which male and female B6C3F1 mice wereexposed to 0, 625, or 1250 ppm (0, 1383, or 2765 mg/m3)butadiene for up to 61 weeks, there were exposure-related increases in the incidences of malignantlymphomas, cardiac haemangiosarcomas (an extremelyrare tumour in B6C3F1 mice), and lung tumours in bothsexes. There were also increased incidences of papillo-mas or carcinomas of the forestomach, hepatocellularadenomas or carcinomas, ovarian granulosa celltumours, acinar cell carcinomas of the mammary gland,brain gliomas, and Zymbal gland carcinomas (the lattertwo tumour types have only rarely been observed in thisstrain of mice at the NTP) in one or both sexes.

Because of the poor survival of mice in the earlierbioassay and to better characterize the exposure–response relationship, the NTP (1993) subsequentlyexposed B6C3F1 mice to lower concentrations (0, 6.25, 20,62.5, 200, or 625 ppm [0, 13.8, 44.2, 138, 442, or 1383mg/m3]) of butadiene for up to 2 years. Survival wasagain decreased in most groups ($20 ppm[$44.2 mg/m3]); at the highest concentration, deathswere principally due to lymphatic lymphomas, whichappeared to arise from the thymus and occurred as earlyas week 23. Non-neoplastic effects observed in exposedmice included a variety of haematological effects, alter-ations in organ weights, bone marrow atrophy andhyperplasia, atrophy of the thymus, atrophy andangiectasis of the ovaries, uterine atrophy, mineralizationof the cardiac endothelium, liver necrosis, and olfactoryepithelial atrophy. There were significant increases in theincidence of tumours at a variety of sites (incidence datapresented in Table 2), including malignant lymphomas(particularly lymphocytic lymphomas), histiocytic sarco-mas, cardiac haemangiosarcomas, Harderian glandadenomas and carcinomas, hepatocellular adenomas andcarcinomas, alveolar/bronchiolar adenomas and carcino-mas, mammary gland adenoacanthomas, carcinomas,and malignant mixed tumours, ovarian granulosa celltumours, and forestomach squamous cell papillomas andcarcinomas, particularly when the incidences wereadjusted for survival. The incidence of alveolar/bronchi-olar adenomas or carcinomas was significantly increasedin females at all concentrations (i.e., $6.25 ppm[$13.8 mg/m3]). Low incidences of uncommon tumours,such as preputial gland carcinomas, Zymbal gland car-cinomas in males, and renal tubule adenomas in bothsexes, were also suspected of being related to exposure.In addition, exposure to butadiene induced malignanttumours at several sites, whereas, in general, tumours atthe same sites in control animals were benign.

The NTP also conducted a “stop-exposure” experi-ment in male B6C3F1 mice designed to investigatewhether tumour induction was associated with theexposure concentration or the duration of exposure.Animals were exposed to 200 ppm (442 mg/m3) for40 weeks or 625 ppm (1383 mg/m3) for 13 weeks (bothequivalent to 8000 ppm-weeks) or to 312 ppm (690 mg/m3)for 52 weeks or 625 ppm (1383 mg/m3) for 26 weeks (bothequivalent to 16 000 ppm-weeks); all groups wereobserved for the remainder of the 2-year study. Again,survival was reduced in all exposed mice, largely due tomalignant neoplasms, with significant increases in theincidences of lymphocytic lymphomas, histiocyticsarcomas, cardiac haemangiosarcomas, Harderian glandadenomas or carcinomas, hepatocellular adenomas orcarcinomas, alveolar/bronchiolar adenomas orcarcinomas, and squamous cell papillomas or carcinomasof the forestomach (even in mice exposed for only 13weeks) (incidence data presented in Table 2). In addition,low incidences of several uncommon tumour types(preputial gland carcinomas, Zymbal gland carcinomas,malignant gliomas and neuroblastomas of the brain,Harderian gland carcinomas, and renal tubule adenomas)were again observed in one or more of the exposedgroups. Concentration may be more important than theduration of exposure in tumour development, as theincidence of malignant lymphomas and squamous cellcarcinomas of the forestomach was greater in the groupsthat had been exposed to 625 ppm (1383 mg/m3) for ashorter period than in those exposed to 200 ppm (442mg/m3) for a longer period (i.e., similar total cumulativeexposure) (NTP, 1993).

Acute exposure of B6C3F1 mice for 2 h to up to10 000 ppm (22 120 mg/m3) butadiene, followed byobservation for 2 years, did not induce an increasedincidence of tumours at any site (Bucher et al., 1993).

Sensitivity to butadiene-induced thymic lympho-ma/leukaemia appears to be enhanced by the presence ofan endogenous ecotropic retrovirus in B6C3F1 mice, asthe incidence of this tumour was greater in male B6C3F1

mice exposed to 1250 ppm (2765 mg/m3) butadiene for 52weeks than in male Swiss mice, which do not express anendogenous retrovirus (57% versus 14%). Exposed miceof both strains had elevated incidences of thymiclymphoma/leukaemia compared with controls, as didB6C3F1 mice exposed to 1250 ppm (2765 mg/m3) for12 weeks and then observed for an additional 40 weeks,although the MuLV env sequence for the retrovirus wasdetected only in tumours of the B6C3F1 mice. Othertumours reported in the mice exposed for 52 weeksincluded haemangiosarcomas of the heart (mainly inB6C3F1 mice) and lung tumours. Neoplasms of theglandular and non-glandular stomach were observed in

Table 2: Incidences of neoplastic lesions in critical carcinogenicity bioassays for butadiene in B6C3F1 mice (NTP, 1993).

Protocol Results Comments

Mice (70 males and 70females per group; 90 malesand 90 females per group atthe highest concentration)were exposed to 0, 6.25, 20,62.5, 200, or 625 ppm (0,13.8, 44.2, 138, 442, or 1383mg/m3) butadiene for 6 h/day,5 days/week, for 103 weeks.Up to 10 mice of each sexfrom each group were killedafter 9 and 15 months ofexposure.

Histopathologicalexamination of acomprehensive range oftissues was carried out onmice in the control and 200and 625 ppm exposure groupskilled after 9 months, on allmice killed at 15 monthsexcept females exposed to6.25 or 20 ppm, and on allmice exposed for 2 years.

Numbers of animals surviving until study termination were 35, 39, 24, 22, 4, and 0 (males) and 37, 33, 24, 11, 0, and 0 (females)at 0, 6.25, 20, 62.5, 200, and 625 ppm, respectively.

Haemangiosarcomas had not previouslybeen observed in 573 male and 558female NTP 2-year historical controls.

Renal tubule adenomas are rare in thisstrain, with a range of 0–1% occurrencein NTP 2-year historical controls.

Carcinomas of the preputial gland arerare in B6C3F1 mice, with none havingbeen observed in NTP historicalcontrols; similarly, sarcomas of thesubcutaneous tissue are uncommon,with the incidence in NTP historicalcontrol females being 2/561.

Zymbal gland neoplasms have not beenobserved in NTP historical controls, norhave carcinomas of the small intestinebeen observed in recent NTP controls.

Exposure to butadiene tended to beassociated with malignant neoplasms inseveral organs, whereas tumours at thesame sites in controls were generallybenign.

Lymphohaematopoietic system Exposure was associated with the development of malignant lymphomas (particularly lymphocytic lymphomas, which occurred asearly as week 23). The incidences were significantly increased in males at 625 ppm (p < 0.001) and females at 200 and 625 ppm(p < 0.001) (although all incidences in the females were within the range of historical control values [8–44%]). Incidences for the 0, 6.25, 20, 62.5, 200, and 625 ppm groups were:males: 4/50, 2/50, 4/50, 6/50, 2/50, and 51/73, or 8, 4, 8, 12, 4, and 70% females: 6/50, 12/50, 11/50, 7/50, 9/50, and 32/80, or 12, 24, 22, 14, 18, and 40%After poly-3 adjustment for survival, the incidences were:males: 9.0, 4.4, 10.0, 15.3, 7.4, and 97.3% females: 13.1, 27.2, 27.5, 20.2, 40.1, and 85.5% Histiocytic sarcomas were significantly increased in both males (p < 0.001) and females (p = 0.002) at 200 ppm, and theincidence of these tumours was marginally higher than that in controls in males at 20, 62.5, and 625 ppm (p = 0.021–0.051) andfemales at 625 ppm (p = 0.038).Incidences: males: 0/50, 0/50, 4/50, 5/50, 7/50, and 4/73, or 0, 0, 8, 10, 14, and 5% females: 3/50, 2/50, 7/50, 4/50, 7/50, and 4/80, or 6, 4, 14, 8, 14, and 5%Adjusted incidences:males: 0, 0, 10.0, 12.9, 24.3, and 50.7% females: 6.5, 4.4, 17.2, 11.8, 34.0, and 35.5%

HeartThe incidences of cardiac haemangiosarcomas were significantly increased compared with controls in males at 62.5 ppm andabove and in females at 200 ppm and above. Incidences:males: 0/50, 0/49, 1/50, 5/48, 20/48, and 4/73, or 0, 0, 2, 10, 42, and 5% females: 0/50, 0/50, 0/50, 1/49, 21/50, and 23/80, or 0, 0, 0, 2, 42, and 29% Adjusted incidences: males: 0, 0, 2.6, 13.5, 64.1, and 52.9% females: 0, 0, 0, 3.1, 71.9, and 83.4%

LungsThere was evidence of increased incidences of alveolar/bronchiolar adenomas or carcinomas compared with controls in males at62.5 ppm and above (p < 0.001) and in females at all concentrations (p < 0.001–0.004).Incidences:males: 21/50, 23/50, 19/50, 31/49, 35/50, and 3/73, or 42, 46, 38, 63, 70, and 4% females: 4/50, 15/50, 19/50, 24/50, 25/50, and 22/78, or 8, 30, 38, 48, 50, and 28%Adjusted incidences:males: 47.5, 49.0, 44.9, 74.2, 87.8, and 45.1% females: 8.8, 33.0, 46.5, 61.1, 81.5, and 82.4%

ForestomachAn increased incidence of forestomach tumours (squamous cell papillomas or carcinomas) was observed in males at 200 and 625ppm (p < 0.001) and females at 62.5 ppm and above (p < 0.001–0.044).Incidences:males: 1/50, 0/50, 0/50, 1/50, 8/50, and 4/73, or 2, 0, 0, 2, 16, and 5% females: 0/50, 0/50, 3/50, 2/50, 4/50, and 22/80, or 0, 0, 6, 4, 8, and 28%Adjusted incidences:males: 2.3, 0, 0, 2.7, 28.7, and 53.5% females: 0, 0, 7.8, 6.1, 22.5, and 82.6%

Table 2 (contd).


OvaryIncreased incidences of malignant and benign granulosa cell tumours were reported in females exposed to 62.5 ppm and above(p < 0.001).Incidences (benign and malignant): 1/49, 0/49, 1/48, 9/50, 8/50, and 6/79, or 2, 0, 2, 18, 16, and 8%Adjusted incidences: 2.3, 0, 2.6, 26.3, 41.1, and 46.5%

Harderian gland The incidence of Harderian gland adenomas and carcinomas was increased in both sexes at 62.5 and 200 ppm (p <0.001–0.016).Incidences:males: 6/50, 7/50, 9/50, 20/50, 31/50 and 6/73, or 12, 14, 18, 40, 62 and 8%females: 8/50, 10/50, 7/50, 15/50, 20/50, and 9/80, or 16, 20, 14, 30, 40, and 11%Adjusted incidences:males: 13.5, 15.2, 22.4, 50.8, 80.6, and 64.2% females: 17.5, 22.7, 17.4, 41.2, 70.9, and 58.0%

Mammary glandThe incidence of mammary gland tumours (adenoacanthomas, carcinomas, and malignant mixed tumours) was increased infemales at 62.5 ppm and above (p < 0.001–0.004). Most of the neoplasms were carcinomas.Incidences: 0/50, 2/50, 4/50, 12/50, 15/50, and 16/80, or 0, 4, 8, 24, 30, and 20% Adjusted incidences: 0, 4.5, 10.2, 32.6, 56.4, and 66.8%

Liver In males, the incidence of hepatocellular neoplasms (adenomas and carcinomas) at 200 ppm was significantly greater than thatin the controls (p = 0.03). The authors also reported increases in hepatocellular neoplasms at 62.5 ppm in females (p = 0.027).Incidences (adenomas and carcinomas):males: 21/50, 23/50, 30/50, 25/48, 33/48, and 5/72, or 42, 46, 60, 52, 69, and 7% females: 15/49, 14/49, 15/50, 19/50, 16/50, and 2/80, or 31, 29, 30, 38, 32, and 3% Adjusted incidences (adenomas and carcinomas): males: 44.6, 48.2, 65.2, 61.6, 85.9, and 61.2% females: 33.3, 30.3, 36.4, 51.4, 64.9, and 21.7%

Other tumoursLow incidences of certain uncommon neoplasms also occurred in exposed mice and were considered to be probably related totreatment. These included preputial gland carcinomas (in 5/50 males at 200 ppm, p < 0.001) and renal tubule adenomas in bothsexes (in 1/50 males at 6.25, 3/48 at 62.5 ppm, and 1/49 at 200 ppm, and in 2/50 females at 200 ppm, compared with none incontrols). Tumours at other sites that “may be related to exposure” included neurofibrosarcomas or sarcomas of the subcutaneous tissue infemales (1/50, 2/50, 3/50, 5/50, 3/50, and 3/80) and Zymbal gland neoplasia (one adenoma in control males, one adenoma andone carcinoma in females at 625 ppm). The investigators were uncertain whether the low incidence of carcinomas of the smallintestine in the treated animals (females: 2 at 6.25 ppm, 1 at 625 ppm, males: 1 each at 6.25, 20, and 62.5 ppm, 2 at 200 ppm,compared with none in controls) was exposure related.

Table 2 (contd).


Mice (males; 50 per group)were exposed to butadiene for6 h/day, 5 days/week, atconcentrations of 200 ppm(442 mg/m3) for 40 weeks(equivalent to a totalexposure of 8000 ppm-weeks),312 ppm (690 mg/m3) for 52weeks (16 000 ppm-weeks), or625 ppm (1383 mg/m3) for 13or 26 weeks (8000 and 16 000ppm-weeks, respectively).After exposure ceased, micewere kept in control chambersuntil 103 weeks andevaluated. Histopathologicalexamination of acomprehensive range oftissues was conducted on allmice.

Lymphohaematopoietic systemThe incidence of malignant lymphomas (the majority of which were lymphocytic lymphomas) was markedly increased in bothgroups exposed to 625 ppm (p < 0.001) and occurred as early as 23 weeks in the 625 ppm (26 weeks) group.Incidences: 4/50, 8/50, 22/50, 8/50, and 33/50, or 8, 16, 44, 16, and 66%, in the control, 200 ppm (40 weeks), 625 ppm (13weeks), 312 ppm (52 weeks), and 625 ppm (26 weeks) groups, respectively.Poly-3 adjusted incidences (adjusted for survival): 9.0, 24.1, 56.1, 35.0, and 87.2%The incidence of histiocytic sarcomas was increased in exposed groups (p < 0.001– 0.036).Incidences: 0/50, 5/50, 2/50, 7/50, and 2/50, or 0, 10, 4, 14, and 4%Adjusted incidences: 0, 16.3, 9.4, 30.6, and 20.4%

Renal tubule adenomas have only rarelybeen observed in NTP historical controls(1/571).

Malignant gliomas and neuroblastomasof the brain rarely developspontaneously in this strain of mice, withnone having been observed in 574 NTPhistorical controls.

HeartThe incidence of cardiac haemangiosarcomas was significantly (p < 0.001) increased in all groups, but particularly in miceexposed to 200 or 312 ppm.Incidences: 0/50, 15/50, 7/50, 33/50, and 13/50, or 0, 30, 14, 66, and 26%Adjusted incidences: 0, 47.1, 30.9, 85.2, and 74.5%

LungsThere was a significant (p < 0.001) increase in the incidence of pulmonary neoplasms (alveolar/bronchiolar adenoma orcarcinoma) in all exposed groups, particularly when the figures were adjusted to take account of mortality.Incidences: 21/50, 36/50, 28/50, 32/50, and 17/50, or 42, 72, 56, 64, and 34%Adjusted incidences: 47.5, 88.6, 89.5, 88.0, and 87.2%

LiverThe incidence of adenomas or carcinomas in the liver was significantly greater in the 200 ppm group (p = 0.004) than in thecontrols and in all exposed groups when adjusted for survival (p < 0.01–0.05).Incidences: 21/50, 33/49, 24/49, 24/50, and 13/50, or 42, 67, 49, 48, and 26%Adjusted incidences: 44.6, 82.4, 80.3, 75.9, and 77.3%

ForestomachThere was a significant (p < 0.001) increase in the incidence of squamous cell papillomas or carcinomas of the forestomach inmice exposed to 312 or 625 ppm (both 13 and 26 weeks).Incidences: 1/50, 3/50, 7/50, 9/50, and 10/50, or 2, 6, 14, 18, and 20%Adjusted incidences: 2.3, 10.2, 28.7, 39.2, and 60.7%

Harderian glandThe incidence of Harderian gland adenomas or carcinomas was significantly (p < 0.001) increased compared with controls in allexposed groups.Incidences: 6/50, 27/50, 23/50, 30/50, and 13/50, or 12, 54, 46, 60, and 26%Adjusted incidences: 13.5, 72.1, 82.0, 88.6, and 76.5%

Other tumoursThe incidences of kidney adenomas were 0/50, 4/48, 1/50, 3/49, and 1/50 in the control, 200 ppm (40 weeks), 625 ppm (13weeks), 312 ppm (52 weeks), and 625 ppm (26 weeks) groups, respectively. The incidence of adenomas or carcinomas of thepreputial gland was significantly (p < 0.001–0.003) increased in the 312 ppm and 625 ppm (13 or 26 weeks) groups, withincidences of 0/50, 1/50, 5/50, 4/50, and 3/50. Malignant gliomas, which were considered to be exposure related, occurred in twomice exposed to 625 ppm for 13 weeks and one exposed to 625 ppm for 26 weeks. Malignant neuroblastomas were observed intwo mice exposed to 625 ppm for 13 weeks. The incidence of adenomas or carcinomas of the Zymbal gland was significantly (p =0.009) increased in mice exposed to 625 ppm for 26 weeks (1/50, 1/50, 0/50, 2/50, and 2/50).


18

the B6C3F1 mice, whereas adenocarcinomas of theHarderian gland and the thyroglossal duct wereobserved in the Swiss mice (Irons et al., 1989).

In the only identified long-term bioassay in rats(Hazleton Laboratories Europe Ltd., 1981a; Owen et al.,1987; Owen & Glaister, 1990), male and female Sprague-Dawley rats were exposed to 0, 1000, or 8000 ppm (0,2212, or 17 696 mg/m3) butadiene for up to 111 weeks. At8000 ppm (17 696 mg/m3), survival was reduced in bothsexes; there were also changes in the relative weights ofa number of organs in males at this concentration, alongwith an increase in the severity of nephrosis of thekidney relative to controls. Relative liver weights wereincreased in all exposed groups, although there were noexposure-related histopathological effects on the liver.At 8000 ppm (17 696 mg/m3), there were increasedincidences of follicular cell adenomas and carcinomas ofthe thyroid gland in females and exocrine adenomas ofthe pancreas in males (with a carcinoma occurring in a ratof either sex) (incidence data presented in Table 3). Infemales, the incidence of benign or malignant mammarygland tumours, along with the incidence of animals withmultiple mammary gland tumours, was increased at both1000 and 8000 ppm (2212 and 17 696 mg/m3). Theincidence of sarcomas of the uterus and carcinomas ofthe Zymbal gland increased significantly with level ofexposure in females; in addition, a Zymbal glandcarcinoma occurred in one male rat at each exposurelevel. The incidence of Leydig cell tumours of the testeswas increased in both groups of exposed males. Theinvestigators suggested that the occurrence of tumoursof the testes and Zymbal gland may have been unrelatedto exposure, as the incidences observed were reportedlysimilar to those in other control rats of the same strain inthe study laboratory, although it is noted that Zymbalgland tumours were noted in the chronic bioassays inmice discussed above.

Both the mono- and diepoxide metabolites (EB andDEB) have induced local tumours at the site of applica-tion in Swiss mice or Sprague-Dawley rats (Van Duurenet al., 1963, 1965, 1966), although available studies areinadequate to evaluate species differences in sensitivity.

It has been hypothesized that the observed greatersensitivity of B6C3F1 mice compared with Sprague-Dawley rats to the induction of thymic lymphoma bybutadiene may be related to differences in the potentialof EB to affect haematopoietic stem cell differentiationobserved in in vitro investigations, as suppression ofclonogenic response was greater in bone marrow cellsfrom C56BL/6 mice than in those from Sprague-Dawleyrats or humans; it was also hypothesized that the sub-population of progenitor cells affected in mice is notpresent in humans (Irons et al., 1995).

8.5 Genotoxicity and related end-points

The genotoxicity of butadiene has been investi-gated in a limited range of in vitro assays and a moreextensive range of in vivo tests. Butadiene was muta-genic in Salmonella typhimurium strains TA1530 andTA1535 in the presence of metabolic activation withrodent or human S9 preparations (de Meester et al., 1978,1980; Arce et al., 1990; NTP, 1993; Araki et al., 1994),although it was generally inactive in strains TA97, TA98,and TA100 with or without exogenous activation undersimilar experimental conditions (Victorin & Ståhlberg,1988; Arce et al., 1990; NTP, 1993). Results of mouselymphoma assays have been conflicting, with anincreased frequency of mutations at the tk locus in onestudy at very high concentrations (i.e., 200 000–800 000ppm [442 400–1 796 600 mg/m3]) in the presence ofmetabolic activation (Sernau et al., 1986), while there wasno convincing activity at concentrations of up to 300 000ppm (663 600 mg/m3) in another study (although theauthors noted that the lack of a positive response mayhave been due to the low solubility of butadiene in theculture medium; NTP, 1993). Butadiene dissolved inethanol induced sister chromatid exchanges in culturedmammalian cells (hamsters and humans) (Sasiadek et al.,1991a, 1991b), while in vitro exposure to gaseousbutadiene did not induce this effect in preparations fromrats, mice, and humans (Arce et al., 1990; Walles et al.,1995).

An overview of the results of available in vivoassays for genotoxicity in germ and somatic cells in miceand rats is presented in Table 4; in general, the data areconsistent with species-specific differences in sensitivityto butadiene-induced genetic damage, likely related tothe quantitative differences in the formation of activemetabolites, although fewer studies have beenconducted in rats. Butadiene induced dominant lethalmutations in two strains of mice (CD-1 and (102/E1 ×C3H/E1)F1) following short-term or subchronic exposureof males to concentrations as low as 500 ppm (1106mg/m3) for 5 days or 65 ppm (144 mg/m3) for 4 weeks;however, exposure to 6250 ppm (13 825 mg/m3) for 6 hdid not induce dominant lethal mutations in CD-1 mice.The results of these studies, which depended upon thetiming of mating relative to exposure, suggested that theinduction of dominant lethal mutations in mice was likelycaused by effects on mature germ cells. In the only simi-lar study in rats identified, there was no evidence ofdominant lethal mutations in Sprague-Dawley ratsexposed to up to 1250 ppm (2765 mg/m3) butadiene for 10weeks.

Short-term exposure to 500 or 1300 ppm (1106 or2876 mg/m3) butadiene also induced an exposure-relatedincrease in the incidence of heritable chromosomal

Table 3: Incidence of neoplastic lesions in critical carcinogenicity bioassays for butadiene in Sprague-Dawley CD rats(Hazleton Laboratories Europe Ltd., 1981a; Owen et al., 1987; Owen & Glaister, 1990).


Rats (110 males and 110 femalesper group) were exposed toconcentrations of 0, 1000, or 8000ppm (0, 2212, or 17 696 mg/m3)butadiene for 6 h/day, 5days/week. Ten rats per sex pergroup were killed at 52 weeks. Thestudy was terminated at 105 weeksin females and 111 weeks inmales. A comprehensive range oftissues from rats at the highconcentration and control rats anda more limited range from rats atthe lower concentration wereexamined microscopically inanimals killed after 52 weeks andat the end of the study.

Survival at study termination was 45, 51, and 32% (males) and 48, 34, and 25% (females) at 0, 1000,and 8000 ppm, respectively (based on interpretation of survival curves in published accounts).

The authors indicated that the incidence ofpancreatic adenomas may be overestimated, due todifficulties in distinguishing between adenomas andhyperplastic foci or nodules in this organ.

The authors noted that the incidence of testiculartumours was similar to that observed in historicalcontrols at Hazleton Laboratories (i.e., 0–6%).

It was stated that the incidences of both uterinesarcomas and Zymbal gland carcinomas were similarto those reported in untreated Sprague-Dawley rats atthe study laboratory and may not have beentreatment related. The authors also indicated thatadditional support for the observed increases intumour incidences not being associated with exposurewas provided by the fact that the majority of theZymbal gland tumours were present in animals killedwithin 76–90 weeks, while none was observed at theend of the study.

Mammary glandThe incidence of total mammary gland tumours (adenomas or carcinomas) was significantly increasedin treated females in both groups (p < 0.01; incidences 50/100, 79/100, and 81/100 in the control,1000, and 8000 ppm groups, respectively). The positive trend was significant (p < 0.001). Theincidence of multiple mammary gland tumours was also increased in exposed females (8/100, 42/100,and 38/100, or 1.38, 3.70, and 3.33 adenomas per adenoma-bearing rat; latter values from Melnick &Huff, 1992).

PancreasThe incidence of pancreatic exocrine adenomas was increased at 8000 ppm in males (p < 0.001)(incidences 3/100, 1/100, and 10/100); exocrine carcinomas also occurred in one male and one femaleexposed to 8000 ppm, compared with none in other groups.

TestesThere was a significant (p < 0.01) exposure-related increase in the incidence of Leydig cell tumours inthe testis (incidences 0/100, 3/100, and 8/100).

Thyroid glandIn females, there were significant exposure-related trends for the occurrence of thyroid follicular celladenomas or carcinomas (0/100, 4/100, and 11/100; p < 0.001) and uterine sarcomas (1/100, 4/100,and 5/100; p < 0.05).

Other tumoursThe incidence of Zymbal gland carcinomas was significantly increased (p < 0.01) at the highestconcentration in females (0/100, 0/100, and 4/100); in males, the incidences were 0/100, 1/100, and1/100.


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Table 4: Overview of genotoxicity of butadiene and its metabolites in rodents.

End-point Mice (strain) Rats (strain) Comments References

BUTADIENE

Germ cells

Dominant lethalmutations

+ (CD-1) ! (Sprague-Dawley) results in mice depended uponduration of exposure and timingof exposure relative to mating;rats were exposed toconcentrations similar to thosethat induced effects in mice

Morrissey et al., 1990;Anderson et al., 1993;Adler et al., 1994, 1998;BIBRA International,1996a, 1996b; Brinkworthet al., 1998

+ ((102/E1 × C3H/E1)F1)

Heritabletranslocations

+ (C3H/E1) NT Adler et al., 1995a, 1998

Other genetic effectson male germ cells(chromosomalaberrations inembryos, DNAdamage, sperm headmorphology,micronuclei)

+ ((102/E1 × C3H/E1)F1) NT Morrissey et al., 1990; Xiao& Tates, 1995; Brinkworthet al., 1998; Pacchierotti etal., 1998a; Tommasi et al.,1998

+ (CD-1)

+ (B6C3F1)

+ (102 × C3H)

Somatic cells

Chromosomalaberrations (bonemarrow)

+ (B6C3F1) NT Irons et al., 1987; Tice etal., 1987; Shelby, 1990;NTP, 1993+ (Swiss)

Sister chromatidexchanges (bonemarrow)

+ (B6C3F1) ! (Sprague-Dawley) rats were exposed to much higherconcentrations than those thatinduced effects in mice

Choy et al., 1986;Cunningham et al., 1986;Tice et al., 1987; Arce etal., 1990; Shelby, 1990;NTP, 1993

Micronuclei (bonemarrow, blood,spleen)

+ (NMRI) ! (Sprague-Dawley) effects in mice were observed atthe lowest concentration tested(i.e., 6.25 ppm); male miceappeared to be more sensitivethan female mice; rats wereexposed to concentrations similarto those that induced effects inmice

Choy et al., 1986;Cunningham et al., 1986;Irons et al., 1986a, 1986b;Tice et al., 1987; Jauhar etal., 1988; Arce et al., 1990;Shelby, 1990; Victorin etal., 1990; NTP, 1993;Przygoda et al., 1993;Adler et al., 1994; Autio etal., 1994; Leavens et al.,1997; Stephanou et al.,1998

+ (B6C3F1) ! (Wistar)

+ (CB6F1)

+((102/E1 × C3H/E1)F1)

+ (NIH Swiss)

hprt– mutations(spleen, thymus)

+ ((102/E1 × C3H/E1)F1) + (F344) mice appeared to be moresensitive than rats

Cochrane & Skopek, 1993,1994b; Tates et al., 1994,1998; Meng et al., 1998,2000

+ (B6C3F1)

! (CD-1)

Specific locusmutations (mousespot test)

+ ((102/E1 × C3H/E1)F1) NT Adler et al., 1994

Transgenic systems(lacZ, lacI)

+ (CD2F1 derived) NT Recio et al., 1992, 1993,1996; Sisk et al., 1994;Recio & Meyer, 1995+ (B6C3F1 derived)

Unscheduled DNAsynthesis (liver)

! (B6C3F1) ! (Sprague-Dawley) Vincent et al., 1986; Arceet al., 1990

DNA–DNA orDNA–protein cross-links (liver)

+/! (B6C3F1) ! (Sprague-Dawley) Jelitto et al., 1989; Ristauet al., 1990; Vangala et al.,1993

DNA binding (liver,lung)

+ (B6C3F1) + (Wistar) levels of adducts were slightlyhigher in mice than in rats

Kreiling et al., 1986; Sorsaet al., 1996b; Koivisto etal., 1997, 1998; Tretyakovaet al., 1998a, 1998b

+ (CB6F1) + (Sprague-Dawley)

+ (F344)


Table 4 (contd).


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DNA strand breaksand other damage(liver, lung, testes)

+ (B6C3F1) + (Sprague-Dawley) results were dependent onanalytical method used; there waslittle quantitative speciesdifference in the degree of strandbreakage

Vangala et al., 1993;Walles et al., 1995;Anderson et al., 1997+ (NMRI)

! (CD-1)

1,2-EPOXY-3-BUTENE (EB)

Germ cells


! ((102/E1 ×C3H/E1)F1)

NT Adler et al., 1997

Other genetic effectson male germ cells(micronuclei)

+ (F1(102 × C3H)) + (Lewis) Lewis rats appeared to be slightlymore sensitive than mice

Xiao & Tates, 1995;Lähdetie et al., 1997;Russo et al., 1997+ (BALB/c) + (Sprague-Dawley)

Somatic cells


+ (C57Bl/6) NT Sharief et al., 1986

Sister chromatidexchanges (spleen)

+ (BALB/c) NT Stephanou et al., 1997

Micronuclei (spleen,blood, bone marrow)

+ (F1(102 × C3H)) + (Lewis) (F1(102 × C3H) mice appeared tobe more sensitive than Lewis rats;CD-1 mice appeared to be moresensitive than Sprague-Dawleyrats

Xiao & Tates, 1995; Adleret al., 1997; Anderson etal., 1997; Lähdetie &Grawé, 1997; Russo et al.,1997; Stephanou et al.,1997

+ (BALB/c) !/+ (Sprague-Dawley)

+ ((102/E1 × C3H/E1)F1)

+ (CD-1)

hprt– mutations(spleen)

+ (B6C3F1) ! (Lewis) Cochrane & Skopek, 1994b;Tates et al., 1998; Meng etal., 1999+ ((102/E1 × C3H/E1)F1) ! (F344)

Transgenic systems(lacI)

! (B6C3F1 derived) + (F344 derived) rats appeared to be moresensitive than mice

Saranko et al., 1998

DNA strand breaksand other damage(bone marrow, testes)

+ (CD-1) +/! (Sprague-Dawley)

damage was observed only inbone marrow cells of rats

Anderson et al., 1997

Unscheduled DNAsynthesis (testes)

! (CD-1) NT Anderson et al., 1997

1,2,3,4-DIEPOXYBUTANE (DEB)

Germ cells


+ ((102/E1 × C3H/E1)F1) NT Adler et al., 1995b

Other genetic effectson male germ cells(chromosomalaberrations inzygotes, micronuclei)

+ ((C57Bl/Cne ×C3H/Cne)F1)

+ (Lewis) Lewis rats appeared to be moresensitive to induction ofmicronuclei than F1(102 × C3H)mice

Adler et al., 1995b; Xiao &Tates, 1995; Lähdetie etal., 1997; Russo et al.,1997+ (F1(102 × C3H)) + (Sprague-Dawley)

+ (BALB/c)

Effects on femalegerm cells(chromosomalaberrations inembryos)

+ (B6C3F1) NT Tiveron et al., 1997

Somatic cells


+ (NMRI) NT positive results were also obtainedin Chinese hamsters, with NMRImice being more sensitive thanhamsters

Walk et al., 1987

Sister chromatidexchanges (bonemarrow, lung, liver)

+ (NMRI) NT positive results were also obtainedin Chinese hamsters, with NMRImice being more sensitive thanhamsters

Conner et al., 1983; Walket al., 1987

+ (Swiss Webster)

+ (BDF1)


Table 4 (contd).


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Micronuclei (spleen,blood, bone marrow)

+ (F1(102 × C3H)) + (Lewis) there was little difference insensitivity between F1(102 × C3H)mice and Lewis rats or betweenCD-1 mice and Sprague-Dawleyrats

Adler et al., 1995b; Xiao &Tates, 1995; Anderson etal., 1997; Lähdetie &Grawé, 1997; Russo et al.,1997; Stephanou et al.,1997

+ (BALB/c) + (Sprague-Dawley)

+ ((102/E1 × C3H/E1)F1)

+ (CD-1)

hprt– mutations(spleen)

+ (B6C3F1) ! (Lewis) F344 rats appeared to be moresensitive than B6C3F1 mice

Cochrane & Skopek, 1994b;Tates et al., 1998; Meng etal., 1999! ((102/E1 ×

C3H/E1)F1)+ (F344)

Transgenic systems(lacI)

! (B6C3F1 derived) ! (F344 derived) Recio et al., 1998

DNA binding + (ICR) NT Mabon et al., 1996

DNA strand breaksand other damage(bone marrow, testes)

+/! (CD-1) +/! (Sprague-Dawley)

damage was noted in bonemarrow cells only

Anderson et al., 1997

Unscheduled DNAsynthesis (testes)

+ (CD-1) NT Anderson et al., 1997

1,2-DIHYDROXY-3,4-EPOXYBUTANE (EBdiol)

Germ cells


! ((102/E1 ×C3H/E1)F1)

NT Adler et al., 1997

Other genetic effectson male germ cells(micronuclei)

NT + (Sprague-Dawley) Lähdetie et al., 1997

Somatic cells

Micronuclei (bonemarrow)

+ ((102/E1 × C3H/E1)F1) + (Sprague-Dawley) Adler et al., 1997; Lähdetie& Grawé, 1997

translocations in mice; an increased incidence ofchromosomal aberrations was also noted in zygotes ofmale mice exposed to $500 ppm ($1106 mg/m3) for5 days. Other butadiene-induced effects observed inmale germ cells of mice include sperm head abnormal-ities, micronuclei in spermatids, and DNA damage(strand breaks and alkaline-labile sites). Investigationsof these end-points in rats have not been identified.

Butadiene was consistently genotoxic in somaticcells of several strains of mice, inducing chromosomalaberrations, sister chromatid exchanges, and micronucleiin numerous assays; micronuclei have been observedfollowing exposure to concentrations as low as 6.25 ppm(13.8 mg/m3) butadiene for 13 weeks or 62.5 ppm(138 mg/m3) for 8 h. Although only few studies wereidentified, these effects were not observed in ratsexposed to much higher concentrations. However, genemutations at the hprt locus have been induced in bothmice and rats, with a four- to sevenfold greater muta-genic potency being determined for mice than for rats.Mutagenic activity was also observed in two transgenicmouse systems and in the mouse spot test. Binding toDNA has been observed in all strains of mice and ratstested; following exposure to butadiene, adducts of bothguanine and adenine with the monoepoxide as well as

the monoepoxide diol metabolites (EB and EBdiol,respectively) have been observed. The degree of adductformation was generally similar in the two species or, insome studies, up to twofold greater in mice than in rats.Similarly, there was little quantitative difference in theamount of butadiene-induced single strand breaks inDNA of mice and rats. DNA–DNA and DNA–proteincross-links were noted in one of two studies in mice, butnot in rats exposed to higher concentrations ofbutadiene.

Metabolites of butadiene have also beenmutagenic and clastogenic in numerous in vitro and invivo assays (see Table 4 for overview of results of invivo assays). EB, DEB, and EBdiol all induced mutationsin bacteria and yeast in the absence of exogenousmetabolic activation (IARC, 1992; NTP, 1993; Thier et al.,1994; Adler et al., 1997); mutagenic activity was alsoobserved for all three metabolites at two foci in humanTK6 lymphoblastoid cells, with DEB being much morepotent (Cochrane & Skopek, 1993, 1994a). Conversely,the monoepoxide was much more potent than thediepoxide in the induction of mutations at the lacItransgene of fibroblasts obtained from a transgenic ratstrain (Saranko & Recio, 1998; Saranko et al., 1998). BothEB and DEB also induced sister chromatid exchanges,


23

chromosomal aberrations, and micronuclei in culturedmammalian (including human) cells (IARC, 1992; Xi et al.,1997). Aneuploidy in chromosomes 12 and X was alsoinduced in human lymphocytes, which is notable in viewof the fact that aneuploidy in these chromosomes iscommonly observed in lymphocytic leukaemias (Xi et al.,1997). In vitro exposure to DEB, but not EB or EBdiol,induced micronuclei in spermatids isolated from rats(Sj`blom & L@hdetie, 1996).

The monoepoxide, diepoxide, and monoepoxidediol metabolites all induced micronuclei in germ cells ofmale mice and rats; in one of these studies, the magni-tude of the effect was greater in Lewis rats than in F1(102× C3H) mice. There were no consistent patterns in therelative potency of the three metabolites. Chromosomalaberrations in zygotes produced by exposed males anddominant lethal mutations were induced by DEB in mice(strains (C57Bl/Cne × C3H/Cne)F1 and (102/E1 × C3H/E1)F1), respectively), whereas EB and EBdiol did notinduce dominant lethal mutations. In the only identifiedinvestigation of the potential effects on female germcells, pre-mating exposure of female B6C3F1 mice to DEBresulted in an increased frequency of chromosomalaberrations in embryos in the absence of ovariantoxicity.

EB, DEB, and EBdiol were also genotoxic insomatic cells (bone marrow, peripheral blood, lung, andspleen), inducing sister chromatid exchanges, chromo-somal aberrations, or micronuclei in several strains ofmice, rats, and hamsters, with little consistent evidenceof interspecies differences in sensitivity; in general, thediepoxide was more potent than the monoepoxide or themonoepoxide diol. Although negative results wereobtained in Lewis rats, both EB and DEB induced anincreased frequency of hprt– mutations in B6C3F1 miceand F344 rats, with rats being more sensitive than mice,which may be related to slower clearance in rats. EBinduced mutations in the bone marrow of lacI transgenicrats, but not in lacI transgenic mice; DEB did not inducelacI mutations in either species. Meng et al. (1999)suggested that the hprt assay is more sensitive to thedetection of large deletions induced by DEB than thelacI transgene assay. DNA damage (strand breaks oralkaline-labile sites) was caused by EB and DEB in thebone marrow of rats and mice, with DEB being lesspotent than EB; the only damage observed in haploidtesticular cells was in mice exposed to EB. It wassuggested that the apparent greater potency of EBcompared with DEB may be due to the bifunctionalalkylating ability of DEB, subsequent induction of DNArepair, and the inability of the alkaline Comet assayemployed to measure cross-links (Anderson et al., 1997).

8.6 Reproductive toxicity

8.6.1 Effects on fertility

Few data on the effects of butadiene on reproduc-tive ability were identified. Exposure to up to 1300 ppm(2876 mg/m3) for 5 days did not affect the reproductiveabilities of male (102/E1 × C3H/E1)F1 mice, based onpercentages of successful pairings with unexposedfemales and unfertilized metaphase I oocytes (Pacchi-erotti et al., 1998a). Similarly, there were no decreasesin mating frequency or pregnancy rate in the dominantlethal studies in mice and rats (Anderson et al., 1993,1998; BIBRA International, 1996a, 1996b; Brinkworth etal., 1998). Documentation of an earlier study in rats,guinea-pigs, and rabbits (Carpenter et al., 1944) is toolimited for evaluation.

The reproductive organs have consistently beentargets of non-neoplastic effects induced by butadienein subchronic and long-term bioassays in B6C3F1 micebut not in Sprague-Dawley rats, although butadiene-induced tumours of the reproductive organs have beenobserved in both species. Ovarian atrophy anddecreased weight were observed in mice exposed to 1000ppm (2212 mg/m3, the only concentration tested) for 13weeks (Bevan et al., 1996). In the 2-year bioassayconducted by the NTP, there was a significant increasein the incidence of ovarian atrophy in females exposedfor up to 2 years to all concentrations tested (i.e., $6.25ppm [$13.8 mg/m3]); both the incidence and the severityof this lesion increased with exposure. Ovarian atrophywas also observed at the interim sacrifices at 9 and 15months at higher concentrations ($200 and $62.5 ppm[$442 and $138 mg/m3], respectively). Atrophied ovariescharacteristically had no evidence of oocytes, follicles,or corpora lutea. Angiectasis and germinal epithelialhyperplasia of the ovaries were reported at $62.5 and$200 ppm ($138 and $442 mg/m3), respectively, afterexposure for 2 years. Uterine atrophy was also noted atconcentrations of 200 ppm (442 mg/m3) or greater.Survival was decreased at $20 ppm ($44.2 mg/m3),principally due to neoplastic lesions at several sites,including the ovaries (Melnick et al., 1990; NTP, 1993).

Effects on the testes, including reduced weight,degeneration, or atrophy, were observed in B6C3F1

mice exposed to concentrations at or above 200 ppm(442 mg/m3) for 2 years or to higher levels for shorterdurations (NTP, 1993; Bevan et al., 1996). Cytotoxiceffects on differentiating spermatogonia were noted in(102/E1 × C3H/E1)F1 mice 21 days after exposure to $130ppm ($288 mg/m3) for 5 days; a decrease in elongatedspermatids was noted in mice exposed to 1300 ppm (2876mg/m3) (Pacchierotti et al., 1998a).


24

No non-neoplastic effects were noted in the repro-ductive organs of male or female Sprague-Dawley ratsexposed to up to 8000 ppm (17 696 mg/m3) butadiene for2 years (Hazleton Laboratories Europe Ltd., 1981a; Owenet al., 1987).

Both the mono- and diepoxide metabolites ofbutadiene induced ovarian toxicity (depletion of smalland growing follicles) and alkylation with macromole-cules in the ovary in B6C3F1 mice repeatedly exposed viaintraperitoneal injection. In contrast, effects in the ovaryin Sprague-Dawley rats were observed only in ratsexposed to the diepoxide at doses higher than those thatwere active in the mice (Doerr et al., 1996). Since theresults of structure–activity studies with 4-vinylcyclo-hexene and several of its analogues, butadiene mono-epoxide and diepoxide, epoxybutane, and isopreneindicated that compounds that form only monoepoxidesdo not induce ovarian toxicity (Doerr et al., 1995), itappears that conversion to the bifunctional diepoxidemay be required for the induction of these effects. Asingle intraperitoneal injection of DEB reduced varioustesticular cell populations and induced morphologicalchanges in the epithelium of the seminiferous tubules inmale B6C3F1 mice (Spano et al., 1996).

8.6.2 Developmental toxicity

Few studies on the potential for butadiene toinduce developmental effects have been identified.There was no evidence of teratogenicity followingexposure of pregnant CD-1 mice to up to 1000 ppm (2212mg/m3) butadiene on days 6 through 15 of gestation,although maternal toxicity (decreased body weight gain)and fetal toxicity (reduced fetal weight and skeletalabnormalities) occurred at 200 ppm (442 mg/m3) andabove, and there was a slight reduction in male fetalbody weight at 40 ppm (88 mg/m3) of questionablebiological significance (Hackett et al., 1987b; Morrisseyet al., 1990). In Sprague-Dawley rats exposed to 8000ppm (17 696 mg/m3) butadiene on days 6 through 15 ofgestation, there was an increased incidence of “major”abnormalities of the skull, spine, sternum, long bones,and ribs. Abnormalities believed to be associated withretarded embryonic growth were observed at 200 and1000 ppm (442 and 2212 mg/m3). Maternal toxicity(decreased body weight gain or loss of body weight)was observed in all exposed groups (HazletonLaboratories Europe Ltd., 1981b, 1982). However, therewas no evidence of developmental toxicity in Sprague-Dawley rats exposed to up to 1000 ppm (2212 mg/m3)butadiene, also on days 6 through 15 of gestation,although maternal toxicity (decreased body weight gain)was noted at the highest concentration (Hackett et al.,1987a; Morrissey et al., 1990).

Although evidence of male-mediatedteratogenicity was observed when male CD-1 miceexposed to 12.5 ppm (27.7 mg/m3) butadiene for 10 weekswere mated with unexposed females (Anderson et al.,1993), there was no increase in malformations when thestudy was repeated at 12.5 and 125 ppm (27.7 and 277mg/m3) (Brinkworth et al., 1998). The authors suggestedthat the discrepant results may be a function of thestatistical significance in the first study being due to thelack of abnormalities in controls (compared with 2.5% inexposed), whereas a low incidence was noted in exposedand control mice in the follow-up study. Similarly, therewere no significant increases in fetal abnormalities inCD-1 mice following paternal exposure to up to 6250 ppm(13 825 mg/m3) butadiene for 6 h (Anderson et al., 1993)or concentrations up to 130 ppm (288 mg/m3) for 4 weeks(BIBRA International, 1996a), although it was noted inthe latter study that some females may have beensacrificed too early for detection of abnormalities. Therewas no evidence of male-mediated teratogenicity inoffspring of male Sprague-Dawley rats exposed for 10weeks to concentrations as high as 1250 ppm (2765mg/m3) butadiene and then mated with unexposedfemales (BIBRA International, 1996b).

8.7 Immunotoxicity

Although the haematopoietic system is a target ofbutadiene-induced toxicity, no effects on immune systemfunction of biological significance were observed in theonly relevant study identified in which B6C3F1 mice wereexposed to 1250 ppm (2765 mg/m3) butadiene for up to 24weeks, although there were depressions in cellularityand plaque-forming cells as well as histopathologicalchanges in the spleen (Thurmond et al., 1986).

9. EFFECTS ON HUMANS

9.1 Clinical studies

Slight irritation of the eyes or respiratory tract wasobserved in volunteers exposed to very high concentra-tions of butadiene (i.e., >2000 ppm [>4400 mg/m3]) in thefew early clinical investigations identified (Larionov etal., 1934; Carpenter et al., 1944); however, these studiesare inadequate for evaluation of the potential effects ofbutadiene in humans, as only subjective symptomsappear to have been evaluated.


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Table 5: Summary of measures of risk for cancers of the lymphohaematopoietic systemin populations occupationally exposed to butadiene.a

Cohort description Cohortsize

Numberof cases Exposure

Risk measureb (95% CI) Comments References

Leukaemia

Styrene-butadienerubber workers

15 649 487

141875

0 ppm-years>0–19 ppm-years20–99 ppm-years100–199 ppm-years$200 ppm-years

SMR = 131 (97–174)RR = 1.0RR = 1.4 (0.4–4.8)RR = 2.3 (0.7–7.9)RR = 2.6 (0.7–10.0)RR = 4.2 (1.0–17.4)

SMR was significantin some subgroups;increasing trend inRRs remained whenadjusted for styrene

Delzell et al.,1995

Butadieneproduction workers

2795996

1874

133

11lowvaried

SMR = 113 (60–193)SMR = 67 (13–195)SMR = 154 (77–275)

no associationbetween qualitativemeasure of cumula-tive exposure andleukaemia risk

Divine &Hartman,1996


364 2 SMR = 123 (15–444) E.M. Ward etal.,1995, 1996

Lymphosarcoma

Styrene-butadienerubber workers

15 649 11 SMR = 80 (40–144) SMRs wereincreased formaintenance workers(O = 8; SMR = 192;95% CI = 83–379)and labourers (O = 3;SMR = 123; 95% CI= 25–359), but not inproduction orlaboratory workers

Delzell et al.,1995


2795 9027

backgroundlowvaried

SMR = 191 (87–364)SMR = 0 (0–591)SMR = 109 (12–395)SMR = 249 (100–513)

no association withduration of employ-ment, based on onlytwo and one cases inthe two higher cate-gories

Divine &Hartman,1996


364 413

<2 years$2 years

SMR = 577(157–1480)SMR = 303SMR = 827 (p < 0.05)

trend with durationof employment whendichotomized at 2years

E.M. Ward etal., 1995,1996

a SMR = standardized mortality ratio; CI = confidence interval; RR = relative risk; O = observed cases.b SMR = observed/expected × 100.

9.2 Epidemiological studies

9.2.1 Cancer

The carcinogenicity of butadiene has beeninvestigated in several populations of workersoccupationally exposed during its manufacture or use.Although most of these studies are limited by thepaucity of historical monitoring data, there is evidencethat occupational exposure to butadiene in the styrene-butadiene rubber industry is associated with excessmortality due to leukaemia and weaker evidence of anassociation with lymphosarcoma1 in butadiene monomerproduction workers. A summary of the measures of risk

for lymphohaematopoietic cancers is presented inTable 5.

In the most recent update of the largest of thecohorts of male monomer workers (n = 2795) at the PortNeches butadiene production facility in Texas, USA(Divine & Hartman, 1996), mortality due to lympho-haematopoietic cancer was significantly elevated(standardized mortality ratio [SMR]2 = 147; 95%confidence interval [CI] = 106–198), due largely to a non-significant increase in the number of deaths due tolymphosarcoma and reticulosarcoma (SMR = 191, 95% CI= 87–364), based on nine cases. However, there was noassociation with duration of employment (SMRs of 261,

1 The terminology for cancers of thelymphohaematopoietic system is that used by authors ofthe individual study accounts.

2 SMRs are presented here in the format used by theauthors; i.e., SMR = observed/expected or SMR =observed/expected × 100.


26

182, and 79, for <5, 5–19, and $20 years of employment,respectively, based on six, two, and one cases). Thegreatest excess was observed in men first employedduring the Second World War (observed cases [O] = 7;SMR = 241; 95% CI = 97–497), during which the greatestconcentrations of butadiene likely occurred, although nodata were presented (Divine et al., 1993). When the entirecohort was subdivided on the basis of a qualitativemeasure of exposure (based on job history information,discussions and reviews of classifications with long-term employees, and recent industrial hygiene surveys),the SMR for lymphosarcoma and reticulosarcoma wasgreatest (SMR = 249; 95% CI = 100–513) in those whosejobs involved “varied” exposure to butadiene (i.e., thegroup with potentially the greatest exposure), althoughthe number of cases in each subgroup was small, andonly one of the seven cases in this group had beenexposed for more than 10 years. There was also a non-significant increase in mortality due to leukaemia in thevaried exposure group (SMR = 154; 95% CI = 77–275;based on 11 cases, 10 of which had been employed lessthan 10 years), although there was no such increase inthe total cohort (SMR = 113; 95% CI = 60–193). However,there was no association between estimates ofcumulative exposure and any form oflymphohaematopoietic cancer.

Mortality due to lymphosarcoma and reticulo-sarcoma was significantly increased in a smaller cohortstudy of 364 workers who had ever been employed inbutadiene production at two Union Carbide plants inWest Virginia, USA, based on four cases (SMR = 5.77;95% CI = 1.57–14.8). The risk was greater in thoseemployed for 2 years or more than in those employed forless than 2 years (SMRs of 8.27 and 3.03, respectively),although the number of cases in each category was verysmall. No significant increase in mortality due toleukaemia or aleukaemia was observed (observed/expected = 2/1.62). However, no monitoring data wereavailable to assess exposure of individuals in the cohort(E.M. Ward et al., 1995, 1996). In the only other study ofmonomer production workers, there were no deaths dueto cancer of the lymphohaematopoietic system;however, the size of the cohort (n = 614) and duration offollow-up were insufficient to detect excess risks oflymphohaematopoietic cancer of less than fivefold(Cowles et al., 1994).

Several studies have been conducted in workersemployed in the manufacture of synthetic rubber inNorth America who were exposed to butadiene as well asto styrene and other substances. The largest and mostcomprehensive study to date involved 15 649 workersemployed at eight styrene-butadiene rubber manufactur-ing facilities in North America (Delzell et al., 1995). Theresults of this study are emphasized here and considered

to supplant those of earlier investigations, as there isconsiderable overlap in the cohort population with theearlier studies (i.e., 14 869 of these subjects had beenemployed at one of the two plants studied previously byMeinhardt et al. [1982] or at seven of the eight plantsinvestigated by Matanoski et al. [1990, 1993] and Santos-Burgoa et al. [1992], although they had been employedfor different time periods [Delzell et al. (1995) includedseveral more years of follow-up] and selected usingdifferent inclusion criteria).1 Estimates of cumulativeexposure and peak exposure frequency were derived forworkers from six of the eight plants based on completework histories for 97% of these employees, informationon processes and plant conditions based on availablerecords, walk-through surveys, and interviews withlong-term employees, plant engineers, and managers andwere compared with monitoring data from surveysconducted from the late 1970s onward.

There was an increase in mortality due to leukae-mia, which was of borderline statistical significance, inthe overall cohort, based on 48 cases (SMR = 131; 95%CI = 97–174); this excess was concentrated in workerswho had had 10 or more years of employment and 20 ormore years since date of hire (O = 29; SMR = 201; 95% CI= 134–288). Similarly, there was a significant increase inmortality due to leukaemia in “ever hourly” workers (i.e.,workers who had ever been paid on an hourly basis)whose jobs were most likely to have involved exposureto butadiene (O = 45; SMR = 143; 95% CI = 104–191),which was again concentrated in workers with longerduration of employment and time since hire and wasgreater in black workers than in white workers in thissubgroup. The SMRs for leukaemia also increased withduration of employment for ever hourly workers. Whenexamined by type of employment, the number of deathsdue to leukaemia was significantly increased inproduction workers (O = 22; SMR = 159; 95% CI =100–241), labourers (O = 16; SMR = 195; 95% CI =112–317; concentrated among black workers), laboratoryworkers (O = 12; SMR = 462; 95% CI = 238–806), andblack workers in other operations (O = 3; SMR = 680;95% CI = 137–1986); no significant increases wereobserved in maintenance workers (O = 13; SMR = 107;95% CI = 57–184). (Although analyses were notpresented and no further information is given, Delzell etal. [1995], in the Discussion section of their report,reported that there was no increase in mortality due toleukaemia in 851 workers in butadiene production areas

1 It is not possible to determine, with any certainty, the

size of the population in these earlier studies that wasnot subsumed in the later investigation by Delzell et al.(1995). One of the small plants of approximately 600workers included in the Matanoski et al. (1990, 1993)cohort was not examined by Delzell et al. (1995).


27

[which would not involve exposure to styrene], withvery low numbers of observed and expected deaths [1versus 2.1, respectively].) As well, the SMRs forlymphosarcoma were non-significantly increased formaintenance workers (O = 8; SMR = 192; 95% CI =83–379) and labourers (O = 3; SMR = 123; 95% CI =25–359), while there was no increase in mortality due tolymphosarcoma in production workers or laboratoryworkers. When individual plants were consideredseparately, there were non-statistically significantincreases in mortality due to leukaemia at most (but notall) plants (SMRs ranged from 72 to 780, excludinggroups in which zero cases were observed and less thanone case was expected); numbers of observed cases oflymphosarcoma were too low to permit meaningfulconclusions with respect to mortality at individualplants.

In regression analyses, mortality due to leukaemiawas observed to increase with cumulative exposure tobutadiene, as relative risk (RR) values for exposurecategories of 0, >0–19, 20–99, 100–199, and $200 ppm-years were 1.0, 1.4, 2.3, 2.6, and 4.2, respectively (forcases in which leukaemia was considered the underlyingcause of death). There was only limited evidence of anassociation with cumulative exposure to peak levels ofbutadiene. The authors also investigated the potentialinfluence of exposure to styrene or benzene on mortalityand determined that the trend for increased risk withincreased cumulative exposure to styrene was lesspronounced, while exposure to benzene was consideredto be too infrequent (few subjects were exposed) and toolow to be a confounding factor. There was noassociation between cumulative exposure to butadieneand non-Hodgkin’s lymphoma in regression analyses.

Based on the results of this study, Delzell et al.(1995) concluded that there was a relationship betweenemployment in the styrene-butadiene industry andleukaemia, with the increased risk of leukaemia beingmost strongly associated with exposure to butadiene orto butadiene and styrene in combination (although theassociation with butadiene remained after controlling forexposure to styrene). Data were insufficient to draw anyfirm conclusions with respect to an association with anyspecific form of leukaemia.

In a subsequent study in which Delzell et al. (1996)attempted to better define exposure of this cohort topeak levels of butadiene, the RR for leukaemia increasedwith increasing average annual number of peaks towhich workers were exposed (RRs of 1.0, 2.3, and 3.1 for0, >0–3288, and >3288 peaks), as well as again withcumulative exposure to butadiene (RRs of 1.0, 1.1, 2.0,2.4, and 4.6 for 0, >0–19, 20–99, 100–199, and $200 ppm-years). Although the analyses were not presented,adjusting for cumulative exposure to styrene apparently

had little influence on the exposure–response relation-ship. Risk of leukaemia also increased with duration ofemployment in areas in which there was “definite” expo-sure to peaks (RRs of 1.0, 2.3, and 2.7 for 0, >0–4, and >5years) and in areas for which elevated SMRs had beennoted in the previous analyses (RRs of 1.0, 1.9, and 3.1for 0, >0–4, and >5 “high SMR-years”). The authorsnoted that it was not possible to distinguish between theroles of estimated peak or cumulative exposure.

The estimates of exposure were further refined forworkers at one of the plants included in the investigationby Delzell et al. (1995) through more extensive researchof historical conditions (Macaluso et al., 1997). Althoughthere was little change in classification of variousworkers as exposed or non-exposed, the revisedestimates of cumulative exposure to butadiene for manyjob groups were generally greater (two- to threefold)than the original; the most substantial increase (by anorder of magnitude) was determined for tasks amongunskilled labourers during the 1950s and 1960s. It wasnot indicated in the report if the rank order of the cumu-lative exposure estimates differed (although it is likelythat it did not; Gerin & Siemiatycki, 1998). There waslittle change in estimated exposure to peak levels ofbutadiene or in cumulative exposure to styrene. Theserevised exposure estimates have not yet been incorpor-ated into cancer mortality analyses.

Sathiakumar et al. (1998) re-examined the mortalityof this cohort based on currently accepted terminologyfor lymphohaematopoietic cancers (other thanleukaemia). There were no significant increases in deathsin the overall cohort due to non-Hodgkin’s lymphoma,Hodgkin’s disease, multiple myeloma, or cancers of otherlymphatic tissue, nor were there any associationsbetween mortality due to these causes and duration ofexposure and year of hire. Similarly, mortality due tothese causes was not associated with any processgroup; however, the authors noted that an associationfor non-Hodgkin’s lymphoma may be obscured by thepossibility that some cases of non-Hodgkin’s lymphomahad transformed to leukaemia, with the latter form ofcancer being recorded on the death certificate.

An association between exposure to butadieneand leukaemia, as well as Hodgkin’s disease, was alsoobserved in a recent independently conducted nestedcase–control study of 58 cases of lymphohaematopoieticcancers from a cohort of styrene-butadiene rubberworkers (from many of the same plants investigated byDelzell et al. [1995]), in which exposure was estimatedbased on analyses of monitoring data obtained in thelast 15–20 years of operation (Matanoski et al., 1997).


28

Irons & Pyatt (1998) noted the concordancebetween the potential for exposure to dimethyldithiocar-bamate (a “stopper” used in the polymerization processin the styrene-butadiene rubber industry) and mortalitydue to leukaemia in various processes. However,although the biological plausibility of a potentialassociation is recognized (since dimethyldithiocarbamateis a potent inhibitor of clonogenic response in humanCD34+ bone marrow cells), it is not possible to draw anyconclusion concerning its potential role in the observedincreases in leukaemia mortality at this time in view ofthe current lack of quantification of exposure levels ofthis substance in the plants examined and the absence ofdata on its leukaemogenic potential.

Other epidemiological studies in populationsoccupationally exposed to butadiene have beenidentified in the literature (e.g., McMichael et al., 1974,1976; Andjelkovich et al., 1976, 1977; Linet et al., 1987;Siemiatycki, 1991; Bond et al., 1992; Downs et al., 1993).However, owing to limitations of these studies(including small numbers of observed and expectedcases of lymphohaematopoietic cancer or lack of expo-sure characterization), they contribute little to evaluationof the association between exposure to butadiene andthese forms of cancer.

Significantly increased mortality due to cancersother than those of the lymphohaematopoietic systemhas not been consistently observed in these studies.

9.2.2 Non-neoplastic effects

Mortality due to all causes was similar to orsignificantly lower than that expected in all of the majorcohorts of workers potentially exposed to butadiene.Although increased mortality due to arteriosclerotic orischaemic heart disease or circulatory disease in generalhas been observed in some subgroups of workers insome of these cohorts (McMichael et al., 1974, 1976;Matanoski et al., 1990; Delzell et al., 1995), the potentialassociation with exposure to butadiene has not beenextensively investigated.

There were no differences in morbidity or varioushaematological parameters between 438 workers exposedto mean concentrations of butadiene ranging up to 10ppm (22 mg/m3) (with a maximum time-weighted averageconcentration of 143 ppm [316 mg/m3]) and 2600unexposed workers at a butadiene production facility inTexas, USA (Cowles et al., 1994). However, Checkoway& Williams (1982) observed changes in haematologicalparameters consistent with bone marrow depression ineight workers exposed to high concentrations ofbutadiene (up to about 53 ppm [117 mg/m3]) whencompared with values for 145 workers exposed to muchlower levels (i.e., <1 ppm [<2.2 mg/m3]).

9.2.3 Genotoxicity

The potential genotoxicity of butadiene hasrecently been investigated in several studies of groupsof workers exposed in the production of butadiene,styrene-butadiene rubber, or polybutadiene rubber.Although the data available to date are not completelyconsistent, they indicate that there is some evidence thatexposure to butadiene induces genetic effects inoccupationally exposed populations and that sensitivityto the induction of these effects is related to geneticpolymorphism for enzymes involved in the metabolism ofbutadiene, most notably those within the glutathione-S-transferase class. The results of several in vitro studiesin human lymphocytes have demonstrated thatsensitivity to DEB-induced sister chromatid exchangesand micronuclei is associated with the presence orabsence of homozygous deletion of the GSTT1 gene,which codes for GST2 (Kelsey et al., 1995; Norppa et al.,1995; Wiencke et al., 1995; Landi et al., 1996; Pelin et al.,1996; Vlachodimitropoulos et al., 1997), for which theprevalence of the null genotype is reported to bebetween 15 and 30% (Nelson et al., 1995; Abdel-Rahmanet al., 1996; Bailey et al., 1998). Similarly, sensitivity tosister chromatid exchanges induced by EB appears to berelated to genotype for GSTM1, which codes for GSTµ(Wiencke & Kelsey, 1993; Uuskula et al., 1995), andpossibly also GSTT1 genotype in GSTM1-nullindividuals (Bernardini et al., 1998). However, there wereno differences in sensitivity to sister chromatidexchanges induced by EBdiol in individuals with andwithout deletions for GSTT1 or GSTM1 (Bernardini et al.,1996).

Although no increased frequencies of sister chro-

matid exchanges, chromosomal aberrations, or micro-nuclei were observed in earlier studies in butadieneproduction workers in Portugal and the Czech Republiccompared with controls (Sorsa et al., 1994, 1996b),positive results for chromosomal aberrations and sisterchromatid exchanges were obtained in the most recentstudy of the Czech workers (Tates et al., 1996; Šrám et al.,1998). When genotype was considered, there was asignificant increase in the frequency of chromosomalaberrations in both exposed and control subjects fromboth plants who were deficient for the GSTT1 gene(Sorsa et al., 1996a).

An increased frequency of hprt– mutants in periph-eral blood lymphocytes has been observed in twostudies of exposed workers at a butadiene productionfacility in Texas, USA (Legator et al., 1993; Ward et al.,1994; Au et al., 1995) and in preliminary results of astudy of styrene-butadiene rubber workers from the


29

same region (J.B. Ward et al., 1996).1 Although analysesby genotype are not yet available, it was noted that thehighest frequency of hprt– variants occurred in anindividual who was GSTT1 null. In contrast to theobservations in the Texan plants, however, no increasein hprt– mutant frequency was observed in workersexposed to similar levels of butadiene at the monomerplant in the Czech Republic (Tates et al., 1996) or in apopulation of polybutadiene rubber workers in China(Hayes et al., 1996) (no information on genotype waspresented). These investigations involved differentanalytical methodologies (autoradiographic versusclonal assays), which may account for the discordancein the results; in addition, differences in occupationalscenarios, exposure levels, age, smoking habits, or otherlifestyle factors may have contributed to thediscrepancy. Current ongoing research (includinggenotyping) may explain the differences in the results.

Decreased DNA repair ability was also observed inperipheral blood lymphocytes of exposed workers at themonomer production and styrene-butadiene rubberfacilities in Texas in both a (-radiation challenge assayand a CAT-Host Cell Reactivation assay (Hallberg et al.,1997).2 However, the difference between exposed and“unexposed” monomer workers in the response to thechallenge assay was no longer significant after ambientlevels in the plant were reduced. Similarly, the effect onDNA repair ability in styrene-butadiene rubber workerswas less when only non-smokers were considered. Thedetection of alkylated DNA (the same adduct as detectedin the liver of mice and rats exposed to butadiene; Jelittoet al., 1989; Koivisto et al., 1997) in the urine of anexposed worker (Peltonen et al., 1993) also providessome evidence of the interaction of butadiene or itsmetabolites with genetic material in humans.

10. EVALUATION OF HEALTH EFFECTS

10.1 Hazard identification

Although the metabolism of butadiene appears tobe qualitatively similar across species, there areextensive data that indicate that the putatively activeepoxide metabolites are formed to a greater degree inmice than in rats. Similarly, although in vivo data arelimited, humans appear to metabolize butadiene to themono- and diepoxide metabolites to a much lesser extentthan mice. However, based on the observed variability inthe formation of adducts of haemoglobin with butadienemetabolites in occupationally exposed human popula-tions, there appears to be interindividual variation inhumans, which is likely related to polymorphism forgenes that code for enzymes involved in the metabolismof butadiene. The weight of evidence for the carcino-genicity, genotoxicity, and non-neoplastic effects ofbutadiene needs to be considered, therefore, in thecontext of these interspecies and interindividualvariations.

10.1.1 Carcinogenicity and genotoxicity

Data supporting the interspecies differences inproduction of active epoxide metabolites are inconcordance with the observed difference in sensitivitybetween mice and rats (at least for the few strainsinvestigated) to butadiene-induced carcinogenicity, inthat the substance appears to be much more potent inmice than in rats. Although butadiene was a multi-sitecarcinogen in both mice and rats at all exposure levelstested (Hazleton Laboratories Europe Ltd., 1981a; NTP,1984, 1993; Irons et al., 1989), the concentrations thatinduced tumours in the only study available in rats weremuch greater than those that were tumorigenic in mice(i.e., $1000 ppm [$2212 mg/m3] versus $6.25 ppm[$13.8 mg/m3]).

Species differences in sensitivity to genetic effectsinduced by butadiene have also been observed.Although butadiene was mutagenic in somatic cells ofboth mice and rats, its mutagenic potency was greater inmice. Other genotoxic end-points (chromosomalaberrations, sister chromatid exchanges, andmicronuclei) were noted in somatic cells of mice but notin those of rats exposed to much higher concentrations.Butadiene was genotoxic in germ cells of male mice inmultiple assays, while negative results were obtained inthe single dominant lethal study in rats. Unlike theobservations with the parent compound, however, thereis little evidence that there are species differences in thesensitivity to genotoxic effects induced by the epoxidemetabolites of butadiene (EB, DEB, and EBdiol),although there was some indication of interstrainvariability. These data suggest that interspeciesdifferences in sensitivity to butadiene-induced

1 Also personal communication (letter dated 17 October1997) from J.B. Ward, Jr., Division of EnvironmentalToxicology, Department of Preventative Medicine andCommunity Health, University of Texas Medical Branch,Galveston, TX, to Health Canada.

2 Personal communication (electronic correspondencedated 15 November 1997) from J.B. Ward, Jr., Division ofEnvironmental Toxicology, Department of PreventativeMedicine and Community Health, University of TexasMedical Branch, Galveston, TX, to Health Canada.


30

genotoxicity are related to quantitative differences in theformation of active metabolites.

There is also limited evidence of the genotoxicityof butadiene in exposed workers; although data are notcompletely consistent, increased frequencies of chromo-somal aberrations, sister chromatid exchanges, and hprt–

mutations and decreased DNA repair capability havebeen reported in some studies of workers in the mono-mer and/or styrene-butadiene rubber manufacturingindustries (Legator et al., 1993; J.B. Ward et al., 1994,1996; Au et al., 1995; Tates et al., 1996; Hallberg et al.,1997; Šrám et al., 1998).1 The discrepancy in the resultsmay be due to the use of different methods for thedetection of mutations or differences in exposure levels.In addition, since sensitivity to induction of geneticeffects by butadiene and its metabolites has been linkedto genotype for glutathione-S-transferase enzymes inseveral in vitro and a few in vivo studies, interpretationof the inconsistent observations in the availabledatabase is complicated by the lack of information ongenotype for most of the small populations examined.

There have been several epidemiological investi-gations of the carcinogenicity of butadiene that serve asa basis for assessment of the weight of evidence forcausality based on traditional criteria. In the most recentcohort study (Delzell et al., 1995), which is also thelargest and most comprehensive investigationconducted to date and that in which exposure was mostextensively characterized, an association betweenexposure to butadiene in the styrene-butadiene rubberindustry and leukaemia was observed (i.e., there was aquantifiable exposure–response relationship). SMRs forleukaemia were elevated for the overall cohort of workersfrom eight plants; the strength of this association wasgenerally greater when specific subgroups with greaterpotential for exposure were considered. In addition, therewas an increase in the RR for leukaemia with increasedcumulative exposure to butadiene in workers from the sixplants for which exposure was best characterized. Theassociation between leukaemia and exposure to buta-diene remained when the potential role of two othersubstances present in the work environment (i.e.,styrene and benzene) was considered. Although furtherrefinement of the estimates of exposure at one of theseplants resulted in increases for several job categories(Macaluso et al., 1997), it is unlikely that these changeswould affect the relative ranking of the categories andanalyses in which exposed workers were compared with“non-exposed” workers (Gerin & Siemiatycki, 1998);

therefore, these results are not inconsistent with theassociation observed by Delzell et al. (1995).

However, no increase in mortality due to leukaemiawas observed in studies of workers involved in theproduction of butadiene monomer who were not con-comitantly exposed to the other substances present inthe styrene-butadiene rubber industry (E.M. Ward et al.,1995, 1996; Divine & Hartman, 1996). Although there wassome evidence of increased mortality due tolymphosarcoma and reticulosarcoma in the subgroup ofworkers potentially exposed to the highestconcentrations of butadiene in the largest of theseinvestigations, there was no association with duration ofemployment or estimated cumulative exposure (based onqualitative ranking of potential for exposure). Althoughmortality due to lymphosarcoma was non-significantlyelevated in some process groups in the styrene-butadiene rubber cohort (Delzell et al., 1995), there wereno consistent patterns (other than for leukaemia), evenwhen currently accepted terminology forlymphohaematopoietic cancers was used (Sathiakumar etal., 1998).

The traditional criterion of consistency for theobserved association between exposure to butadieneand leukaemia is fulfilled, at least in part, in that similarexcesses were observed among plants in the large cohortstudy of styrene-butadiene rubber workers (Delzell et al.,1995); i.e., there is internal consistency. A similarexposure–response was also noted in an independentnested case–control study of mostly the samepopulation in which different exposure assessmentmethodology was employed (Matanoski et al., 1997).Observation of external consistency with results of othercohort studies of styrene-butadiene rubber workers islargely precluded, in view of the scope of the largeepidemiological cohort study that included a largeproportion of all of the styrene-butadiene rubber workersin North America. Indeed, it is difficult to envisageadditional studies in this occupational group that wouldcontribute meaningfully to weight of evidence forconsistency of the observed association.

One criterion for causality of observedassociations in epidemiological studies, namelycoherence, may not have been adequately fulfilled, inview of the difference in the specific form oflymphohaematopoietic cancer in excess in availableinvestigations for the two principal types of populationsof workers studied. Indeed, increases in lymphosarcomaand reticulosarcoma have been observed in monomerproduction workers, whereas increases in leukaemia havebeen observed in styrene-butadiene rubber workers.Although it is plausible that this difference may berelated to variation in the extent of information availablefor characterization of exposure or to the nature ofexposures in the two industries, this has not beensystematically investigated. There is also the possibility

1 Also personal communications (letter dated 17 October1997 and electronic correspondence dated 15 November1997) from J.B. Ward, Jr., Division of EnvironmentalToxicology, Department of Preventative Medicine andCommunity Health, University of Texas Medical Branch,Galveston, TX, to Health Canada.


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of misclassification of cause of death on deathcertificates (although Sathiakumar et al. [1998] did notobserve an association with forms of lymphohaemato-poietic cancer other than leukaemia in the large cohort ofstyrene-butadiene rubber workers when causes of deathwere examined using current terminology). The potentialfor transformation of one form of lymphohaematopoieticcancer to another (e.g., non-Hodgkin’s lymphoma toleukaemia) has also been noted (Sathiakumar et al., 1998).In addition, available data for the large study of styrene-butadiene rubber workers were insufficient to determineif butadiene was causally associated with a specific formof leukaemia. Moreover, it is noteworthy that thesedifferent tumours observed in styrene-butadiene rubberworkers and monomer production workers are of thesame organ system, and perhaps even share the samepluripotential stem cell.

An association between exposure to butadieneand the induction of leukaemia is also biologicallyplausible. The haematopoietic system is a target forbutadiene-induced effects in rodents (i.e., lymphocyticlymphomas [NTP, 1993], cytogenetic effects in bonemarrow [Cunningham et al., 1986; Irons et al., 1986a,1987; Tice et al., 1987; NTP, 1993; Leavens et al., 1997],and suppression of stem cell differentiation [Irons et al.,1996]). Aneuploidy, which is believed to be associatedwith leukaemia in humans, has been induced in humanlymphocytes exposed in vitro to the mono- and diepox-ide metabolites of butadiene (Vlachodimitropoulos et al.,1997; Xi et al., 1997). Moreover, the presence of relevantmetabolizing enzymes in progenitor cells believed to beimportant targets for the induction of leukaemia inhumans (i.e., CD34+ cells) has been demonstrated instudies of the metabolism of benzene (a documentedhuman leukaemogen) (Schattenberg et al., 1994; Ross etal., 1996) (although exposure of human CD34+ cells to EBat “physiologically relevant concentrations” did not altercytokine-induced clonogenic response, an early changefrequently observed in the development of leukaemia;Irons et al., 1996).1 Therefore, available data also supportthe biological plausibility of an association betweenexposure to butadiene and leukaemia observed inhumans, although the active metabolite has not beenidentified.

Therefore, although not completely convincing intheir own right, the available epidemiological studies ofthe association between leukaemia and exposure tobutadiene in occupationally exposed human populationsfulfil several of the traditional criteria for causality,including strength of association (RR of 4.2 in thehighest exposure group [based on five cases], whichwould be considered moderately strong), quantifiableexposure–response relationship, temporal relationship(the critical investigation [i.e., Delzell et al., 1995] is ahistorical cohort study), biological plausibility, and, tosome degree, consistency, although the criterion forcoherence is not fully satisfied.

Assessment of the weight of evidence forcarcinogenicity in human populations should not,however, be considered in isolation from the extensivesupporting data on carcinogenicity, genotoxicity, andinter- and intraspecies variations in metabolism andresponse. The association between exposure tobutadiene and development of cancer is supported bylimited evidence of genetic damage in exposed workers,as well as the wealth of evidence that butadiene iscarcinogenic and/or genotoxic in all species ofexperimental animals tested (mice, rats, and hamsters),inducing a wide range of tumours and genetic damage atrelatively low concentrations in mice (i.e., within thesame order of magnitude as current occupational healthlimits). Moreover, while there are quantitative differencesin the potency of the substance to induce tumours invarious species, likely related to observed quantitativedifferences in metabolism, there are indications ofconsiderable interindividual variations in the metabolismof butadiene in the human population, consistent withexpectations for a complex metabolic pathway.

The observation of an association betweenexposure in the occupational environment and leukaemiathat fulfils several of the traditional criteria for causalityof associations observed in epidemiological studies, aswell as supporting limited data on genotoxicity in humanpopulations and the well documented carcinogenicityand genotoxicity at relatively low concentrations in somespecies of experimental animals, provides weight ofevidence that butadiene is carcinogenic in humans.

Although relevant data in humans are limited, the

results of in vivo studies in experimental animals indicatethat butadiene induces mutations in somatic cells andmale germ cells as well as male-mediated heritableclastogenic damage. While most of the studies havebeen conducted in mice, rats appear to be less sensitiveto these effects, which is consistent with speciesdifferences in metabolism. However, in view of the likelyconsiderable heterogeneity in the metabolism ofbutadiene in human populations, butadiene may be ahuman somatic and germ cell genotoxicant.

1 Also personal communication (correspondence dated30 March 1998) from R.D. Irons, University of ColoradoHealth Sciences Center, Denver, CO, to Health Canada.


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The available data on effects of butadiene otherthan carcinogenicity or genotoxicity are limited. Basedon the limited data available, species differences in theability of butadiene to induce other non-neoplasticeffects again appear to be consistent with variations inmetabolism of butadiene to active metabolites. However,butadiene is of low acute toxicity in both rats and mice,in contrast to its ability to induce cancer and geneticdamage at relatively low concentrations in mice.

Haematological effects suggestive of macrocyticanaemia have been consistently observed in mice (twostrains) following short-term, subchronic, or chronicexposure to butadiene at concentrations similar to orlower than those that induced general toxicity (asindicated by decreased body weight gain and increasedorgan weights) (Irons et al., 1986a, 1986b; NTP, 1993;Bevan et al., 1996). For example, changes in haemato-logical parameters were noted in mice exposed to$62.5 ppm ($138 mg/m3) butadiene for 9 months orlonger in the NTP bioassay. Butadiene also inducedeffects on bone marrow (including atrophy, decreasedcellularity, regeneration, and alterations in stem celldevelopment) in mice (Irons et al., 1986a, 1986b;Leiderman et al., 1986; NTP, 1993), although availabledata are inadequate to assess the potential effects onimmune system function. While effects on the blood andbone marrow have not been reported in rats in recentinvestigations (including the only identified chronicbioassay; Hazleton Laboratories Europe Ltd., 1981a), thedatabase is considerably more limited. In addition, thelack of observation of haematotoxicity in rats may againreflect the species differences in metabolism. Althoughthe available epidemiological studies are too limited toassess the haematotoxicity in humans, available datasupport the haematopoietic system being a critical targetfor butadiene-induced toxicity, since thelymphohaematopoietic system is a target for butadiene-induced leukaemia in humans. However, it has not beenestablished if the non-neoplastic effects observed inanimals may be preliminary to, or associated with, thedevelopment of lymphohaematopoietic cancers.

The reproductive organs are also critical targets ofbutadiene-induced non-neoplastic effects in mice.Ovarian atrophy, the severity and incidence of whichincreased with concentration or duration of exposure,was observed at all concentrations (i.e., $6.25 ppm[$13.8 mg/m3]) in the chronic bioassay conducted by theNTP (1993); in all exposure groups, the level of degen-eration at 2 years, characterized by lack of oocytes,follicles, or corpora lutea, was incompatible withreproductive capacity. Although recent re-examinationof some of the tissue samples indicated that the atrophyobserved in the ovaries may be related to senile

changes,1 it may be that butadiene is exacerbating thesechanges. It should be noted, though, that the incidenceof these lesions was increased as early as 9 months(although the slides from these interim sacrifices havenot been re-examined). That butadiene is causallyassociated with these lesions is also difficult to dismisson the basis of currently available data, in view of theconsistency with the results of other studies, includingthe earlier NTP (1984) bioassay and a subchronic studyat higher concentrations (Bevan et al., 1996) in whichsuch lesions were also observed, the presence of a cleardose–response relationship, and biological plausibility.Based on the observation of depletion of ovarianfollicles and alkylation with ovarian macromolecules inmice following intraperitoneal administration of themonoepoxide or diepoxide metabolite and in ratsadministered the diepoxide (Doerr et al., 1995), it ispossible that the ovarian toxicity is mediated throughgeneration of the active epoxide metabolites.

Testicular atrophy was noted only in male miceexposed to concentrations greater than those thatinduced effects in females (NTP, 1993). Consistent withmetabolic differences, butadiene did not induce ovarianor testicular toxicity in the limited number of availablestudies in rats, although, as noted above, the diepoxidemetabolite was ovotoxic in both species (Doerr et al.,1995, 1996).

Although available data are limited, there is noconclusive evidence that butadiene is teratogenic inmice or rats following maternal or paternal exposure orthat it induces significant fetal toxicity at concentrationsbelow those that are maternally toxic.

Available epidemiological data are inadequate forevaluation of potential reproductive or developmentaltoxicity; in fact, none of the identified analytical studieswas conducted in women. However, in view of the quali-tative similarities in the metabolism of butadiene in mice,rats, and humans and the likely variation across the gen-eral population associated with genetic polymorphismfor the relevant enzymes, and on the basis of theobserved ovarian toxicity in butadiene-exposed mice,butadiene may be a reproductive toxicant in humans,although additional work to clarify the relevance of theseobserved effects is clearly desirable.

Available data on other systemic or organ-specificeffects are inadequate to determine if such effects mightbe considered critical.

1 Personal communication (electronic correspondencedated 26 June 1998) from B. Davis, National Institute forEnvironmental Health and Safety, National ToxicologyProgram, Research Triangle Park, NC, to Health Canada.


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10.2 Exposure–response assessment andcriteria for setting tolerableconcentrations or guidance values

As per the approach adopted for several genotoxiccarcinogens, measures of the potency of a substance toinduce effects for which there is believed to be nothreshold may be used to establish guideline values forenvironmental media. Since air is the principal route ofexposure to butadiene in the general environment(available data indicate that other routes contributenegligibly), quantitation of exposure–response forcancer as well as non-cancer effects is limited toexposure by inhalation.

In order to eliminate the uncertainty associatedwith extrapolation from animal species, quantitativemeasures of carcinogenic potency (i.e., tumorigenicconcentrations, or TCs)1 have been developed on thebasis of available epidemiological data. This is based onthe conclusion that the weight of evidence for anassociation between butadiene and leukaemia satisfiesseveral of the traditional criteria for causality inepidemiological studies. However, uncertainties in theexposure estimates for the critical cohort of workers aswell as confounding or effect-modifying aspects thatcould impact on quantitative estimates of risk arerecognized. In view of these factors and to serve as abasis for comparison, quantitative measures of cancerpotency have also been developed on the basis ofresults of long-term bioassays in rats and mice, withthose in mice being considered justifiably conservative,considering the likely heterogeneity in metabolictransformation of butadiene in humans. (See discussionof relevance of specific tumour types in animals tohumans in section 10.2.1.2.)

In addition to inducing tumours at multiple sites inexperimental animals, butadiene is also genotoxic insomatic and germ cells and induces reproductive andhaematological effects in animals. As a measure of

exposure–response for non-cancer effects, where con-sidered appropriate, benchmark concentrations2 havebeen calculated on the basis of data from long-termstudies in mice.

Several physiologically based pharmacokinetic(PBPK) models have been developed as a basis forreducing uncertainty in interspecies extrapolations forbutadiene by various groups of investigators. However,none of the models currently available has adequatelyaccounted for the distribution of metabolites in thecompartments included; the principal researchers in thisfield have concluded that there are likely more factorsinvolved in butadiene metabolism than have beenincluded in the models developed to date (Csanády etal., 1996; Sweeney et al., 1997). In addition, none of themodels has included the formation of EBdiol, a puta-tively active metabolite that is believed to be importantin humans, since it has been observed to bind to haemo-globin to a greater degree than EB in workers exposed tobutadiene. Nor has bone marrow been incorporated as acompartment, although it appears to be a target site ofbutadiene-induced toxic effects. Moreover, none of thePBPK models has been validated in humans. For thesereasons, therefore, such models have not been used toquantitatively account for interspecies variations inmetabolism in the quantitation of exposure–response forcritical end-points based on studies in experimentalanimals presented here.

In addition, owing to its relatively slow metabo-lism, butadiene achieves a steady state during prolongedinhalation exposure. Exposures of the same concentra-tion and duration would be expected to result in equiva-lent toxicity across species, and interspecies scaling toaccount for variations in inhalation rate to body weightratios or body surface areas between humans andanimals is not considered necessary.

10.2.1 Carcinogenicity

10.2.1.1 Epidemiological data

In only one epidemiological investigation of theassociation between butadiene and leukaemia have dataon exposure of the study population been sufficientlycharacterized to permit quantitation of exposure–response (Delzell et al., 1995). The Delzell et al. (1995)study also presents results for the largest cohort studied

1 The potency estimate for carcinogenicity is determined

by calculating the dose or concentration associated withan increase in cancer incidence or mortality of anappropriate percentage. When based on toxicologicaldata from studies in experimental animals, a 5% increaseis generally chosen, as these values usually lie within orclose to the observable range (i.e., a TC05 is calculated).When epidemiological data form the basis for derivationof a tumorigenic concentration, the percent increaseselected is that which falls within the area of theexposure–response curve that represents the majority ofthe observable data; this is often less than 5%. In thecase of butadiene, the carcinogenic potency calculatedon the basis of modelling of epidemiological data (asdescribed herein) was considered to be best defined as a1% increase in mortality due to leukaemia (i.e., a TC01).

2 Similar to tumorigenic concentrations (TC05s),benchmark concentrations for non-cancer effects (orBMC05s), when based on data in experimental animals,represent the dose or concentration associated with a5% increase in the incidence of an effect compared withcontrols.


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to date (including subjects from eight plants, six ofwhich were included in the exposure–responseanalyses); it is also considered to subsume theobservations of mortality in workers at these plantsreported previously by other researchers (i.e., Meinhardtet al., 1982; Matanoski et al., 1990, 1993; Santos-Burgoaet al., 1992), because of the considerable overlap in thecohort definition. The exposure assessment of studysubjects was of extremely high quality, being verythorough and based on research of plant recordsconcerning work histories, processes and localemissions, and consultation with staff from each plant,and is, therefore, considered appropriate forquantification of exposure–response (limited industrialhygiene monitoring data were also available, althoughused primarily for comparison with estimatedconcentrations). For comparison with estimates basedon the data from the cohort study, carcinogenic potencywas also calculated on the basis of the results of thecase–control study nested within essentially the samepopulation of workers (Matanoski et al., 1997), althoughdata available in the published report were too limited topermit detailed analysis here.

A detailed description of the methods employedand the assumptions made in the derivation of tumori-genic potency estimates (i.e., TC01, or the concentrationassociated with a 1% increase in mortality due toleukaemia based on the observations in Delzell et al.[1995]) is presented in Appendix 4. Although severalmathematical models were applied, the choice of modelhad little impact on the resulting TC01, as there was onlya threefold variation in the range of values. However, theestimated TC01 in which confidence is highest (i.e., thatfor which the model provided the best fit) is 1.7 mg/m3.The TC01 calculated on the basis of the data of Matan-oski et al. (1997) was only slightly lower than this value.

10.2.1.2 Data from studies in experimental animals

As described in section 10.1.1, butadiene inducedan increase in the incidence of tumours at multiple sitesin both B6C3F1 mice (liver, lung, Harderian gland, mam-mary gland, ovaries, forestomach, Zymbal gland, andkidney, along with malignant lymphomas, histiocyticsarcomas, and cardiac haemangiosarcomas) andSprague-Dawley rats (mammary gland, thyroid gland,uterus, Zymbal gland, pancreas, and testes). Asdiscussed above, consistent with the species differencesin metabolism, mice were much more sensitive tobutadiene-induced cancer than were rats for the strainsinvestigated. Based on data available (i.e., evidence fromgenotoxicity studies that butadiene and its metabolitesare active in both species), this difference in sensitivityis quantitative rather than qualitative and is related tothe greater amounts of putatively active metabolites

formed in mice compared with rats. In addition, thedifferent profiles of tumours observed in the two speciesmay be related to differential roles of the epoxidemetabolites in the induction of the various tumours; i.e.,the diepoxide may be more critical to tumour induction inmice than is EB (since it was reported recently thatformation of DEB increased with level of exposure tobutadiene in mice but not in rats; Thornton-Manning etal., 1998), while the monoepoxide or monoepoxide diolmay be more important in rats.

The relevance for extrapolation to humans ofexposure–response for some of the types of tumoursobserved in rodents has been questioned. For example,Irons et al. (1989) hypothesized that the thymic lympho-ma/leukaemia induced in B6C3F1 mice may be related tothe presence of an endogenous ecotropic retrovirus, as amuch lower incidence was observed in Swiss mice thatdo not possess this retrovirus (although the incidencewas significantly elevated compared with controls).Therefore, although the haematopoietic system is atarget for the induction of cancer by butadiene inhumans, the observed exposure–response relationshipfor this end-point is not considered appropriate forquantitative extrapolation to humans — on the basis thatthis retrovirus is not present in humans and its presencein B6C3F1 mice renders this strain quite susceptible toinduction of lymphoma — although the relevant infor-mation is included for comparative purposes.

It has also been suggested that the tumoursobserved in the study in rats (i.e., mammary gland,thyroid gland, pancreas, uterus, and testes) and some ofthe tumours induced in mice (i.e., ovaries and mammarygland) may be mediated through effects on theendocrine system. Indeed, tumours at these sites areoften associated with disruption of hormonally mediatedfunctions. In addition, non-neoplastic or pre-neoplasticeffects, including atrophy, degeneration, andhyperplasia, have also been observed in mice exposedsubchronically to butadiene. However, the mechanismby which butadiene induces tumours at these sites hasnot yet been adequately investigated; i.e., it has notbeen established whether these tumours are induced viaa mechanism for which there may be a threshold ofexposure (e.g., through induction of hormonallymediated effects), although the possibility is recognized.In addition, the results of in vivo genotoxicity assaysindicate that butadiene or its metabolites induce geneticeffects in the reproductive organs of multiple strains ofmice.

Based on these considerations, estimates of car-cinogenic potency were calculated on the basis of themalignant lymphomas, histiocytic sarcomas, cardiachaemangiosarcomas, alveolar/bronchiolar adenomas or


35

carcinomas, hepatocellular adenomas or carcinomas,squamous cell papillomas or carcinomas of the fore-stomach, adenomas or carcinomas of the Harderiangland, granulosa cell tumours of the ovaries, andadenoacanthomas, carcinomas, or malignant mixedtumours of the mammary gland observed in B6C3F1 micein the chronic bioassay conducted by the NTP (1993)and the mammary gland tumours, pancreatic exocrineadenomas, Leydig cell tumours, Zymbal glandcarcinomas, thyroid follicular cell adenomas or carci-nomas, and uterine sarcomas in Sprague-Dawley ratsreported by Hazleton Laboratories Europe Ltd. (1981a). Itis noted that the characterization of exposure–responseis much better in the study in mice (which involved fiveclosely spaced exposure levels) than in the bioassay inrats (in which only two more widely spaced exposurelevels were used, the higher of which was likely abovethe level of metabolic saturation). (Although there werealso increased incidences of tumours at several sites inB6C3F1 mice in the “stop-exposure” study conducted bythe NTP [1993], only TC05s determined on the basis ofthe 2-year study were included, as the latter studyprovides better information for characterization ofexposure–response in mice following long-term exposure[i.e., more exposure levels for up to 2 years].)

The methods employed for development of esti-mates of tumorigenic potency (i.e., the concentrationsassociated with a 5% increased incidence of tumours, orTC05s) based on these data are described in Appendix 4.TC05s based on observations in mice ranged from2.3 mg/m3 (95% lower confidence limit [LCL] = 1.7 mg/m3)for Harderian gland tumours in males to 99 mg/m3 (95%LCL = 23 mg/m3) for malignant lymphomas in males. Forrats, calculated TC05s ranged from 6.7 mg/m3 (95% LCL =4.7 mg/m3) to 4872 mg/m3 (95% LCL = 766 mg/m3) fortumours of the mammary gland and Zymbal gland infemales, respectively.

Although not presented here, modelling of theincidence of micronucleated polychromatic erythrocytesin B6C3F1 mice exposed to butadiene for up to 15 monthsin the NTP bioassay resulted in a benchmarkconcentration (BMC05) that was very similar to the lowerend of the range of estimates of tumorigenic potency.


There have been recent attempts to quantitativelyestimate risk of heritable genetic damage in humansbased on a parallelogram approach and data on male-mediated heritable translocations and bone marrowmicronuclei in mice and chromosomal aberrations inlymphocytes of exposed workers (Pacchierotti et al.,1998b). In view, however, of the reported ovarianatrophy due to reduction of primordial follicles (to adegree that would preclude reproduction) following

chronic exposure of mice to concentrations of butadieneconsiderably lower than those associated with adverseeffects on the testes, investigation of the response offemale germ cells in mice to butadiene is desirable, sincethis may well be the most sensitive end-point for devel-opment of quantitative estimates of heritable damage.(Determination of putatively toxic metabolites in theovaries of butadiene-exposed female mice would also beinformative.) For this reason, quantitation of exposure–response for heritable genetic damage is not presentedhere. However, in view of the apparent greatersensitivity of the reproductive organs in female mice, abenchmark concentration was derived for non-neoplasticeffects in the ovary, which is considerably moreprotective than that for male-mediated heritable damagedeveloped by Pacchierotti et al. (1998b). (Although therelative role of butadiene in the induction of theobserved atrophy in mice in the NTP study is unclear, asdiscussed in section 10.1.2, information currentlyavailable is not considered a sufficient basis upon whichto dismiss this end-point as being inappropriate forquantification of exposure–response. However, thisuncertainty should be kept in mind in the interpretationor application of the BMC05s derived below.)

Ovarian atrophy was observed in both long-termNTP (1984, 1993) bioassays in mice and a subchronicstudy (Bevan et al., 1996). Although limited, availabledata indicate that rats are less sensitive to induction ofthis effect, which may, again, be a consequence of inter-species variations in metabolism. Therefore, althoughadditional research into the etiology of the observedovarian atrophy in mice would be desirable, the datafrom the later NTP study are considered most appropri-ate for characterization of exposure–response (i.e.,development of a BMC05). In this investigation, theincidence of atrophy of the ovaries was significantlyincreased in an exposure-related manner at all concen-trations tested (i.e., $6.25 ppm [$13.8 mg/m3]). Theseverity of this effect also increased with exposure.

The derivation of the BMC05 for ovarian atrophy ispresented in detail in Appendix 4. Because the expo-sure–response curve plateaus at the higher exposurelevels, the two highest exposure groups were omittedfrom the calculations. The resulting BMC05 for ovarianatrophy in mice was determined to be 0.57 mg/m3 (95%LCL = 0.44 mg/m3), when all degrees of severity wereconsidered. If only lesions of moderate or markedseverity were considered, the resulting BMC05 would beabout fivefold higher.

Haematotoxicity is considered to be a critical effectassociated with exposure to butadiene. Although thehaematopoietic system appears to be a target forbutadiene-induced cancer in humans, available data onthe potential non-neoplastic effects on this system are


36

inadequate for quantitation of exposure–response.However, since statistically significant changes wereobserved in mice only at concentrations greater thanthose that induced other toxic effects, and sincebenchmark concentrations derived for effects on theblood are greater than those for these other effects,quantitation of the exposure–response forhaematological effects has not been presented here.

10.3 Sample exposure and riskcharacterization

10.3.1 Sample exposure characterization

The principal source of environmental exposure tobutadiene is air. Although few data were identifiedregarding levels in drinking-water and food, due to itsphysical/chemical properties (e.g., vapour pressure andpartition coefficients) and environmental release patterns(i.e., principally atmospheric emissions), intake ofbutadiene in these media is expected to be negligible incomparison with that in air.

As an example of population exposure characteri-zation, estimates are presented on the basis of dataavailable for Canada. Based on concentrations measuredin outdoor air in several rural, suburban, and urbanlocations across Canada1 (see section 6.1.1), 95% of thegeneral population can be expected to be exposed toaverage concentrations of up to 1.0 µg/m3. However,since levels are generally greater in highly urbanizedareas, estimated “reasonable worst-case exposure” isexpected to be up to 1.3 µg/m3 (95th percentile). In areasinfluenced by industrial point sources, exposure couldbe as high as 6.4 µg/m3, based on the 95th percentile ofconcentrations measured near a source in Ontario(MOEE, 1995).

Individuals may also be exposed to butadiene forshort durations while at self-service gasoline fillingstations or in parking garages; however, these intakesare still much less than average daily intakes for thegeneral population from inhalation of backgroundconcentrations in outdoor and indoor air.

Although available Canadian data indicate thatbutadiene is detected with greater frequency in indoorair than in outdoor air, there are insufficient data to

characterize the distributions of concentrations ofbutadiene in various indoor environments. In general,butadiene is detected more frequently and at higherconcentrations in indoor environments contaminated byETS than in areas where smoking does not occur. Non-smokers who spend a considerable proportion of theirtime in indoor environments where ETS is present can beexposed to concentrations of butadiene that are an orderof magnitude higher than the average levels in theoutdoor air. Tobacco use (e.g., 20 cigarettes per day) canincrease the daily intake of butadiene by smokers by fivetimes over the daily intake by non-smokers in ETS-contaminated indoor locations. The daily intake ofbutadiene by smokers can be 100 times greater than thedaily intake of non-smokers who are not exposed to ETS.

10.3.2 Sample risk characterization

Butadiene is released to air from both industrialpoint sources and more dispersive, non-point sources,the latter due to its production primarily duringincomplete combustion. Based on estimates derivedusing monitoring data from Canada, intake for thegeneral population is primarily from air, with intake fromother media likely being negligible in comparison. Thefocus of the human health risk characterization is,therefore, the general population exposed in outdoor andindoor air in the general environment and those exposedthrough air in the vicinity of industrial point sources.

For compounds such as butadiene, where data aresufficient to support a plausible mode of action forinduction of tumours by direct interaction with geneticmaterial, estimates of exposure are compared withquantitative estimates of cancer potency to characterizerisk.

Tumorigenic concentrations were calculated on thebasis of data from both epidemiological studies andinvestigations in experimental animals. For the criticalepidemiological investigation (Delzell et al., 1995), a TC01

(i.e., the concentration associated with a 1% increase inmortality due to leukaemia) was considered theappropriate measure of carcinogenic potency, since themajority of the observable data fell within this range.Although four different mathematical models wereconsidered, the TC01 generated by the model with thebest fit was 1.7 mg/m3.

Quantitative estimates of carcinogenic potencyderived on the basis of data in experimental animals werecalculated as TC05s (i.e., the concentration associatedwith a 5% increase in tumour incidence). Based on the 2-year bioassay in mice (NTP, 1993), TC05s ranged from 2.3mg/m3 (95% LCL = 1.7 mg/m3) to 99 mg/m3 (95% LCL = 23mg/m3). The TC05s derived on the basis of the morelimited study in rats (Hazleton Laboratories Europe Ltd.,

1 Unpublished data on butadiene levels in Canada fromNational Air Pollution Surveillance program, provided byT. Dann, River Road Environmental Technology Centre,Environment Canada, Ottawa, Ontario, to CommercialChemicals Evaluation Branch, Environment Canada, Hull,Quebec, April 1997.


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1981a) ranged from 6.7 mg/m3 (95% LCL = 4.7 mg/m3) to4872 mg/m3 (95% LCL = 766 mg/m3).

The values derived on the basis of studies inhumans are preferred as the basis for comparison withestimates of exposure to characterize risk. While there area number of uncertainties in the use of the epidemi-ological data for both hazard evaluation and exposure–response analyses (section 10.4), these are likely far lessthan uncertainties associated with interspecies extrapo-lation. Moreover, estimated potency for humans issimilar to that developed on the basis of the cancerbioassays in experimental animals. (Indeed, although inan area of the exposure–response curve where data weremore sparse, it is noteworthy that TC05s calculated onthe basis of epidemiological data [as opposed to theTC01s presented above] are within the range of valuesderived from the studies in rodents.)

Based on the sample exposure scenarios presentedabove (section 10.3.1), 95% of the population in Canadais exposed to concentrations of butadiene in outdoor airof 1.0 µg/m3 or less. For the proportion of the generalpopulation that is regularly exposed to higher concentra-tions of butadiene in urban areas (i.e., the “reasonableworst-case scenario”), the 95th percentile of the distrib-ution of concentrations is 1.3 µg/m3. In the only area ofCanada identified as having an industrial point source,the 95th percentile of the distribution of concentrationsis 6.4 µg/m3.

The margins between carcinogenic potency andestimated exposure for the general population (includingambient and reasonable worst case) and those in thevicinity of a point source are presented in Table 6.Equivalent low-dose risk estimates are also presented inthis table.

In view of the relative potency of butadiene toinduce some non-cancer effects, these end-points arealso important in risk characterization. As presentedabove, a benchmark concentration (BMC05) of 0.57 mg/m3

(95% LCL = 0.44 mg/m3) was derived on the basis of datafor the incidence of ovarian atrophy of all severities (i.e.,female reproductive toxicity) in mice exposed tobutadiene for up to 2 years (NTP, 1993). And while thereis uncertainty about the relevance of the ovarian atrophyobserved in mice for humans (section 10.4), the BMC05 isslightly less than the lower end of the range of estimatesof cancer potency based on the incidence of tumours inthe same study in mice, as well as the TC05 for cancerbased on the epidemiological data. The mode ofinduction of ovarian atrophy is unknown. However, if itis (reasonably) assumed that the mode of action isrelated to that by which tumours are induced (i.e., directinteraction with genetic material), risk to human healthfor reproductive effects may be characterized in the samemanner as presented for cancer. Therefore, estimates of

the margin between the BMC05 for ovarian toxicity and asample exposure characterization are presented in Table7. It should be noted, though, that even if the mode ofinduction of ovarian atrophy does not involve directinteraction with genetic material, the margin betweenexposure and effect level (i.e., for which a tolerableconcentration is normally developed) is still small — i.e.,exposure levels in Canada are 90–570 times lower thanthe benchmark concentration, as presented in Table 7.

10.4 Uncertainties and degree ofconfidence in human health hazardcharacterization and sample riskcharacterization

There is a high degree of certainty that butadieneis being released to ambient air in Canada in significantamounts in vehicular exhaust. There is a moderatedegree of certainty that exhaust emissions of butadieneare lower in well maintained vehicles equipped withcatalytic converters than in older non-equipped vehicles,and that evaporative emissions during refuelling andvehicle operation contribute less to concentrations ofbutadiene in ambient air than do emissions in vehicularexhaust.

There is a moderate degree of certainty that buta-diene is not being released to the Canadian environmentin significant amounts from industrial activities inCanada, as only a single major point source (i.e., inSarnia, Ontario) of discharge to the atmosphere has beenidentified. Although there is some uncertainty that theavailable measurements of butadiene in samples takenover a few days in the vicinity of this source are repre-sentative of population exposure over the long term,since the samples were taken at distances of up to a fewkilometres from the source, there is a moderate degree ofcertainty that a segment of the population would beexposed to the measured concentrations. There is a highdegree of certainty that populations in rural areas areexposed to lower concentrations of butadiene in ambientair than are communities in more densely populatedareas.

Available data on concentrations of butadiene inambient air in Canada are quite extensive. A large pro-portion of the numerous samples from several samplingsites across the country contained concentrations ofbutadiene above the level of detection. Therefore, thereis a high degree of certainty in the estimations of expo-sure to butadiene via ambient air.


38

Table 6: Comparison of estimates of carcinogenic potency with exposure levels.

Exposure Potency (TC01 or TC05)

Margin betweenpotency andexposure

Equivalent low-dose riskestimate

1.0 µg/m3 (95th percentilefor all sites in Canada)

1.7 mg/m3 (TC01 for leukaemia in humans) 1700 5.9 × 10–6

2.3 mg/m3 (TC05 for most sensitive tumoursite in mice [Harderian gland])

2300 22 × 10–6

1.7 mg/m3 (95% LCL of TC05 for mostsensitive tumour site in mice)

1700 29 × 10–6

6.7 mg/m3 (TC05 for most sensitive tumoursite in rats [mammary gland])

6700 7.5 × 10–6

4.7 mg/m3 (95% LCL of TC05 for mostsensitive tumour site in rats)

4700 11 × 10–6

1.3 µg/m3 (95th percentilefor reasonable worst-casescenario)



1800 28 × 10–6


1300 38 × 10–6


5200 9.6 × 10–6


3600 14 × 10–6

6.4 µg/m3 (95th percentilefor area affected byindustrial point source)



360 14 × 10–5


270 19 × 10–5


1000 5.0 × 10–5


730 6.8 × 10–5

Table 7: Comparison of estimates of potency for non-cancer effects with exposure levels.

Exposure Potency (BMC05)

Margin betweeneffect level and

exposure

1.0 µg/m3 (95th percentile for all sitesin Canada)

0.57 mg/m3 (BMC05 for ovarian atrophy in mice) 570

0.44 mg/m3 (95% LCL of BMC05 for ovarian atrophy in mice) 440

1.3 µg/m3 (95th percentile forreasonable worst-case scenario)



6.4 µg/m3 (95th percentile for areaaffected by industrial point source)



The most limiting aspect of the exposure assess-ment is the lack of sufficient data on the concentrationsof butadiene in indoor air. This is an important short-coming, since humans spend significantly more time inindoor environments than outdoors. In the absence ofindoor sources, it is reasonably certain that concentra-tions of butadiene in indoor environments are similar tothe concentrations in the local ambient air.

Higher concentrations of butadiene have beenmeasured in indoor air where ETS was known to bepresent. However, the data on concentrations of buta-diene in ETS-contaminated indoor air are highly variableand are not sufficient to reasonably define the range ofmean concentrations. Nevertheless, there is a highdegree of certainty that non-smokers spending aconsiderable proportion of their time in indoorenvironments where ETS is present are exposed to


39

higher concentrations of butadiene than are non-smokers who are not exposed to ETS. There is a highdegree of certainty that smokers are exposed to higherconcentrations of butadiene and have significantlyhigher daily intakes than do non-smokers.

There is somewhat less certainty that butadienemonomer is not released in detectable amounts fromconsumer products (e.g., synthetic materials) incorpor-ating this compound in their production. Although theremay be contributions to indoor concentrations of buta-diene from certain cooking activities, the data are notsufficient to identify specific sources or activities or toidentify a range of emissions of butadiene duringcooking.

Although data on levels of butadiene in foodstuffsare scarce, based on the physical and chemicalproperties of the substance and the fact that it isreleased primarily to ambient air (where it is likely toremain without partitioning to other media), there is areasonable degree of certainty that food does notrepresent a major source of exposure. Similarly, althoughthe database for concentrations of butadiene in drinking-water is limited, there is a reasonable degree of certaintythat drinking-water is not an important source ofexposure for the general public in Canada, based on thevolatility and release patterns of the compound.

There is some degree of uncertainty that theweight of epidemiological evidence for the associationbetween butadiene and leukaemia satisfies criteria forcausality. In particular, the need for coherence isseemingly not addressed, since the observed increase inmortality due to leukaemia in styrene-butadiene rubberworkers was not observed in the cohorts of monomerworkers (although there was some evidence of anassociation with other forms of lymphohaematopoieticcancer, particularly in short-term workers). This may berelated to the nature of exposure to both butadiene andother substances in these two industries. However, inview of the overwhelming evidence of carcinogenicityand genotoxicity in experimental animals, available infor-mation on species differences in sensitivity likely beingrelated to differences in metabolism, and the potential forconsiderable interindividual variability in metabolism toputatively toxic metabolites in the human population,along with the limited evidence of genotoxicity in occu-pationally exposed populations, there is a high degree ofconfidence that butadiene is likely to be carcinogenic inhumans. Based on the extensive database on the geno-toxicity of butadiene and its principal metabolites both invitro and in vivo in both somatic and germ cells, confi-dence that butadiene induces tumours (and possiblyother effects) through direct interaction with geneticmaterial is high.

Although the assessment of the exposure of thecritical cohort of workers is likely one of the mostcomprehensive published to date, there is also uncer-tainty in the estimates of carcinogenic potency derivedon the basis of this study, due primarily to the fact thatthe estimates of exposure are based on few actual histori-cal monitoring data.1 For example, when the exposure ofworkers at one plant was re-examined, there were 2- to 3-fold changes in the estimates for several job groups(with a 10-fold increase for one job group). In addition,with the exception of incorporating exposure to styreneas a stratification variable in the analyses, potentialinteractions between various occupational exposurescould not be taken into account in the derivation of thecarcinogenic potency based on the observations in thiscohort. It has also been demonstrated that genetic poly-morphism for several of the enzymes involved in metabo-lism of butadiene affects sensitivity to toxic effectsinduced by the substance. Also, since information ongenotype for the relevant enzymes was not available forthis large cohort and only a small amount of informationon the distribution in the general population has beenidentified, it is not possible to determine how represen-tative the study cohort is of the genetic susceptibility tobutadiene of the general public.

With respect to the quantitation of exposure–response and derivation of potency estimates based onthe epidemiological data, the inability of any of themodels to consistently predict leukaemia rates in thevalidation study contributes to additional uncertainty. Inaddition, the small number of leukaemia cases beingmodelled contributes to model instability. However, thefact that the range of potency estimates for the fourmodels is narrow (i.e., 1.4–4.3 mg/m3) increases theconfidence in the calculated potencies. There is alsosome uncertainty associated with the fact that thepotency estimates were based only on cases in whichleukaemia was considered the underlying cause of death,rather than all cases, which could result in an underesti-mation of leukaemogenic potency.

In view of the likely variability in metabolism ofbutadiene across the human population related togenetic polymorphism for relevant enzymes, estimates ofcarcinogenic potency as well as benchmarkconcentrations for non-cancer effects based on studiesin mice are considered justifiably conservative. However,

1 Although it has not been possible to quantitativelycharacterize uncertainty regarding these estimates ofexposure and the impact of this uncertainty upon theestimates of carcinogenic potency, data being collectedcurrently may permit a more quantitative characterizationin future (personal communication [correspondencedated 20 March 1998] from J. Lynch, Consultant,Rumson, NJ, to Health Canada).


40

because of the high mortality in the study in mice inwhich exposure–response could best be characterizedand the limitations in the study in rats (high mortality atthe higher of only two widely spaced exposure levels),there is a moderate degree of uncertainty in estimates ofcarcinogenic potency derived on the basis ofinvestigations in experimental animals. In addition, sincethe available PBPK models were considered inadequate,the potency estimates developed here are based on aninhaled concentrations exposure metric; i.e., none of theavailable information on species differences inmetabolism is taken into account. It is noteworthy that ifthe calculated margins between exposure andcarcinogenic potency presented above that serve tocharacterize risk were derived on the basis of the 95%LCLs of the TC05s for tumours in mice, the values woulddiffer by only 1.4- to 3.3-fold (i.e., within the same orderof magnitude) from those calculated on the basis of thepoint estimates; similarly, use of the 95% LCLs of theTC05s for tumours in rats would result in a 1.1- to 6.4-folddifference in the margins between exposure and potency.In addition, although confidence in the use of thepotency estimate for lymphomas in mice is low, due tothe inherent sensitivity associated with the presence ofan endogenous retrovirus, it is noteworthy that the TC05

for this tumour would not be limiting, as it falls within therange of values determined for cancers at other sites.Also, it should be noted that, although these marginsand measures of risk presented above were based oncomparison of the 95th percentile of the exposure datafor each scenario, use of the median concentration (i.e.,the 50th percentile) and either the point estimates ofcarcinogenicity or the associated 95% LCLs would resultin a 5-fold difference in the resulting values for thegeneral population and a 10-fold difference in values forthose in an area influenced by a point source.

There is uncertainty about the relevance of theovarian atrophy observed in mice to humans, based onlack of data on the relative role of butadiene in theetiology of these lesions. Although the observed effectsin the 2-year bioassay may have been related to senilechanges, possibly exacerbated by butadiene, atrophy ofthe ovaries was detected in these mice as early as9 months, and there is consistent evidence in otherchronic and subchronic studies that the ovaries are tar-gets of toxic effects induced by butadiene or its epoxidemetabolites. As a result of this uncertainty, quantitativemeasures of dose–response developed on the basis ofovarian atrophy must necessarily be interpreted withcaution. In addition, the BMC05 presented above wasbased on inclusion of atrophy of all severities, including“minimal” severity, the biological significance of whichis unclear. If only lesions of moderate or marked severityare considered, the resulting BMC05 and hence the cal-culated margin between exposure and effect level andrisk estimates would differ by about fivefold. (Use of the95% LCLs of the BMC05s for atrophy of all severities or

of only moderate or marked severity would result in onlya 1.5- or 3-fold difference in the measure of risk.) How-ever, in view of the weight of evidence of causality forthe association between butadiene and these effects inmice and the relatively low value for the measure ofdose–response compared with that for other types ofeffects, additional investigation in this area is deemed tobe of high priority.

12. PREVIOUS EVALUATIONS BYINTERNATIONAL BODIES

An IARC Working Group that convened in 1998has classified 1,3-butadiene as probably carcinogenicto humans (Group 2A) based on limited evidence ofcarcinogenicity in humans and sufficient evidence ofcarcinogenicity for 1,3-butadiene and 1,2:3,4-diepoxy-butane in experimental animals (IARC, 1999).


41

REFERENCES

Abdel-Rahman SZ, El-Zein RA, Anwar WA, Au WW (1996) A multiplexPCR procedure for polymorphic analysis of GSTM1 and GSTT1 genesin population studies. Cancer letters, 107:229–233.

Adler I-D, Cao J, Filser JG, Gassner P, Kessler W, Kliesch U,Neuhauser-Klaus A, Nusse M (1994) Mutagenicity of 1,3-butadieneinhalation in somatic and germinal cells of mice. Mutation research,309:307–314.

Adler I-D, Filser JG, Gassner P, Kessler W, Schoneich J, Schriever-Schwemmer G (1995a) Heritable translocations induced by inhalationexposure of male mice to 1,3-butadiene. Mutation research,347:121–127.

Adler I-D, Kliesch U, Tiveron C, Pacchierotti R (1995b) Clastogenicityof diepoxybutane in bone marrow cells and male germ cells in mice.Mutagenesis, 10(6):535–541.

Adler I-D, Kliesch U, Nylund L, Peltonen K (1997) In vitro and in vivo

mutagenicity of the butadiene metabolites butadiene diolepoxide,butadiene monoepoxide and diepoxybutane. Mutagenesis,12(5):339–345.

Adler I-D, Filser J, Gonda H, Schriever-Schwemmer G (1998) Doseresponse study for 1,3-butadiene-induced dominant lethal mutations andheritable translocations in germ cells of male mice. Mutation research,397:85–92.

Albrecht OE, Filser JG, Neumann H-G (1993) Biological monitoring of1,3-butadiene: species differences in haemoglobin binding in rat andmouse. In: Sorsa M, Peltonen K, Vainio H, Hemminki K, eds. Butadiene

and styrene: Assessment of health hazards. Lyon, International Agencyfor Research on Cancer, pp. 135–142 (IARC Scientific Publications No.127).

Altshuller AP, Lonneman WA, Sutterfield FD, Kopezynski SL (1971)Hydrocarbon composition of the atmosphere of the Los Angeles basin1967. Environmental science and technology, 5:1009–1016.

Anderson D, Edwards AJ, Brinkworth MH (1993) Male-mediated F1effects in mice exposed to 1,3-butadiene. In: Sorsa M, Peltonen K,Vainio H, Hemminki K, eds. Butadiene and styrene: Assessment of

health hazards. Lyon, International Agency for Research on Cancer, pp.171–181 (IARC Scientific Publications No. 127).

Anderson D, Dobrzynka MM, Jackson I, Yu T-W, Brinkworth MH (1997)Somatic and germ cell effects in rats and mice after treatment with 1,3-butadiene and its metabolites, 1,2-epoxybutene and 1,2,3,4-diepoxybutane. Mutation research, 391:233–242.

Anderson D, Hughes JA, Edwards AJ, Brinkworth MH (1998) Acomparison of male-mediated effects in rats and mice exposed to 1,3-butadiene. Mutation research, 397:77–84.

Andjelkovich D, Taulbee J, Symons M (1976) Mortality experience of acohort of rubber workers, 1964–1973. Journal of occupational medicine,18:387–394.

Andjelkovich D, Taulbee J, Symons M, Williams T (1977) Mortality ofrubber workers with reference to work experience. Journal of

occupational medicine, 19:397–405.

Araki A, Noguchi T, Kato F, Matsushima T (1994) Improved method formutagenicity testing of gaseous compounds by using a gas samplingbag. Mutation research, 307:335–344.

Arce GT, Vincent DR, Cunningham MJ, Choy WN, Sarrif AM (1990) Invitro and in vivo genotoxicity of 1,3-butadiene and metabolites.Environmental health perspectives, 86:75–78.

Atkinson R, Aschmann SM, Winer AM, Pitts JN (1984) Kinetics of thegas-phase reactions of NO3 radicals with a series of dialkenes,cycloalkenes, and monoterpenes at 291 K. Environmental science and

technology, 18:370–375.

Atkinson R, Arey J, Aschmann SM, Long WD, Tuazon EC, Winer AM(1990) Lifetimes and fates of toxic air contaminants in California’s

atmosphere. Final report. Prepared by Statewide Air Pollution ResearchCenter, University of California, Riverside, CA, for California AirResources Board, California Environmental Protection Agency, March(Contract No. A732-107).

Au WW, Bechtold WE, Whorton EB Jr, Legator MS (1995) Chromosomeaberrations and response to (-ray challenge in lymphocytes of workersexposed to 1,3-butadiene. Mutation research, 334:125–130.

Autio K, Renzi L, Catalan J, Albrecht OE, Sorsa M (1994) Induction ofmicronuclei in peripheral blood and bone marrow erythrocytes of ratsand mice exposed to 1,3-butadiene by inhalation. Mutation research,309:315–320.

Bailer AJ, Portier CJ (1988) Effects of treatment-induced mortality andtumor-induced mortality on tests for carcinogenicity in small samples.Biometrics, 44:417–431.

Bailey LR, Roodi N, Verrier CS, Yee CJ, Dupont WD, Parl FF (1998)Breast cancer and CYP1A1, GSTM1, and GSTT1 polymorphisms:evidence of a lack of association in Caucasians and AfricanAmericans. Cancer research, 58:65–70.

Barrefors G (1996) Air pollutants in road tunnels. Science of the total

environment, 189/190:431–435.

Batinka IB (1966) Maximum permissible concentrations of divinylvapors in the air of work areas. Gigiena i Sanitariya, 31:18–22.

Bechtold WE, Strunk MR, Chang I-Y, Ward JB Jr, Henderson RF (1994)Species differences in urinary butadiene metabolites: comparisons ofmetabolite ratios between mice, rats, and humans. Toxicology and

applied pharmacology, 127:44–49.

Bechtold WE, Strunk MR, Thornton-Manning JR, Henderson RF (1995)Analysis of butadiene, butadiene monoxide, and butadiene dioxide inblood by gas chromatography/mass spectrometry. Chemical research in

toxicology, 8:182–187.

Becker KH, Biehl HM, Bruckmann P, Fink EH, Führ F, Klöpffer W,Zellner R, Zetzsch C (1984) Methods of the ecotoxicological evaluation

of chemicals. Photochemical degradation in the gas phase. Vol. 6. OH

reaction rate constants and tropospheric lifetimes of selected

environmental chemicals. Report 1980–1983. KernforschungsanlageJülich GmbH. Projektträger Umweltchemikalien.

Bell RW, Chapman RE, Kruschel BD, Spencer MJ, Smith KV, Lusis MA(1991) The 1990 Toronto personal exposure pilot (PEP) study. Preparedfor Atmospheric Research and Special Programs Section, AirResources Branch, Ontario Ministry of the Environment, Toronto,Ontario (ARB-207-90).

Bell RW, Chapman RE, Kruschel BD, Spencer MJ (1993) Windsor Air

Quality Study. Personal exposure survey results. Toronto, Ontario,Ontario Ministry of Environment and Energy, Science and TechnologyBranch.

Bernardini S, Pelin K, Peltonen K, Jarventaus H, Hirvonen A, Neagu C,Sorsa M, Norppa H (1996) Induction of sister chromatid exchange by3,4-epoxybutane-1,2-diol in cultured human lymphocytes of differentGSST1 and GSTM1 genotype. Mutation research, 361:121–127.

Bernardini S, Hirvonen A, Pelin K, Norppa H (1998) Induction of sisterchromatid exchange by 1,2-epoxy-3-butene in cultured humanlymphocytes: influence of GSTT1 genotype. Carcinogenesis,19(2):377–380.

Bevan C, Stadler JC, Elliott GS, Frame SR, Baldwin JK, Leung H-W,Moran E, Panepinto AS (1996) Subchronic toxicity of 4-vinylcyclohexene in rats and mice by inhalation exposure. Fundamental

and applied toxicology, 32:1–10.

BIBRA International (1996a) The detection of dominant lethal mutations

and foetal malformations and chromosome damage in the offspring of male

mice treated sub-chronically with butadiene by inhalation — second

study. Carshalton, Surrey (Report No. 1542/1).


42

BIBRA International (1996b) The detection of dominant lethal mutations

and foetal malformations in the offspring of male rats treated sub-

chronically with 1,3-butadiene by inhalation. Carshalton, Surrey (ReportNo. 1542/2).

Bond GG, Bodner KM, Olsen GW, Cook RR (1992) Mortality amongworkers engaged in the development or manufacture of styrene-basedproducts — an update. Scandinavian journal of work, environment and

health, 18:145–154.

Bond JA, Dahl AR, Henderson RF, Dutcher JS, Mauderly JL, BirnbaumLS (1986) Species differences in the disposition of inhaled butadiene.Toxicology and applied pharmacology, 84:617– 627.

Boogaard PJ, Bond JA (1996) The role of hydrolysis in thedetoxification of 1,2:3,4-diepoxybutane by human, rat and mouse liverand lung in vitro. Toxicology and applied pharmacology, 141:617–627.

Boogaard PJ, Sumner SC-J, Turner MJ, Bond JA (1996a) Hepatic andpulmonary glutathione conjugation of 1,2:3,4-diepoxybutane in human,rat, and mouse in vitro. Toxicology, 113:297–299.

Boogaard PJ, Sumner SC-J, Bond JA (1996b) Glutathione conjugation of1,2:3,4-diepoxybutane in human liver and rat and mouse liver and lungin vitro. Toxicology and applied pharmacology, 136:307–316.

Brinkworth MH, Anderson D, Hughes JA, Jackson LI, Yu T-W,Hieschlag E (1998) Genetic effects of 1,3-butadiene on the mousetestis. Mutation research, 397:67–75.

Brunnemann KD, Dagan MR, Cox JE, Hoffmann D (1989)Determination of benzene, toluene and 1,3-butadiene in cigarette smokeby GC-MSD. Experimental pathology, 37:108–113.

Bucher JR, Melnick RL, Hildebrandt PK (1993) Lack of carcinogenicityin mice exposed once to high concentrations of 1,3-butadiene. Journal

of the National Cancer Institute, 85(22):1866–1867.

Camford Information Services (1995) CPI product profiles. Don Mills,Ontario.

CARB (1992) Technical support document. Proposed identification of

1,3-butadiene as a toxic air contaminant. California Air ResourcesBoard, Stationary Source Division.

Carpenter CP, Shaffer CB, Weil CS, Smyth HF Jr (1944) Studies onthe inhalation of 1:3-butadiene; with a comparison of its narcotic effectwith benzol, toluol, and styrene, and a note on the elimination of styreneby the human. Journal of industrial hygiene and toxicology, 26:69–78.

CEH-SRI International (1994) Butadiene. Chemical EconomicsHandbook-SRI International (CEH Marketing Research Report).

Checkoway H, Williams TM (1982) A hematology survey of workers ata styrene-butadiene synthetic rubber manufacturing plant. American

Industrial Hygiene Association journal , 43:164–169.

Choy WN, Vlachos DA, Cunningham MJ, Arce GT, Sarrif AM (1986)Genotoxicity of 1,3-butadiene. Induction of bone marrow micronuclei inB6C3F1 mice and Sprague-Dawley rats in vivo. Environmental

mutagenesis, 8(Suppl. 6):18.

Cochrane JE, Skopek TR (1993) Mutagenicity of 1,3-butadiene and itsepoxide metabolites in human TK6 cells and in splenic T cells isolatedfrom exposed B6C3F1 mice. In: Sorsa M, Peltonen K, Vainio H,Hemminki K, eds. Butadiene and styrene: Assessment of health

hazards. Lyon, International Agency for Research on Cancer, pp.195–204 (IARC Scientific Publications No. 127).

Cochrane JE, Skopek TR (1994a) Mutagenicity of butadiene and itsepoxide metabolites: I. Mutagenic potential of 1,2-epoxybutene, 1,2,3,4-diepoxybutane and 3,4-epoxy-1,2-butanediol in cultured humanlymphoblasts. Carcinogenesis, 15:713–717.

Cochrane JE, Skopek TR (1994b) Mutagenicity of butadiene and itsepoxide metabolites: II. Mutational spectra of butadiene, 1,2-

epoxybutene and diepoxybutane at the hprt locus in splenic T cells fromexposed B6C3F1 mice. Carcinogenesis, 15:719–723.

Conner MK, Luo JE, Gutierrez de Gotera O (1983) Induction and rapidrepair of sister-chromatid exchanges in multiple murine tissues in vivo

by diepoxybutane. Mutation research, 108:251–263.

Conor Pacific Environmental (1998) A report on multimedia exposures to

selected PSL2 substances. Prepared by Conor Pacific Environmental(formerly Bovar Environmental) and Maxxam Analytics Inc. for HealthCanada, Ottawa, Ontario (Project No. 741-6705; Contract # DSS FileNo. 025SS.H4078-6-C574).

Cowles SR, Tsai SP, Snyder PJ, Ross CE (1994) Mortality, morbidity,and haematological results from a cohort of long term workers involvedin 1,3-butadiene monomer production. Occupational and environmental

medicine, 51:323–329.

CPPI (1997) Technical dossier — 1,3-Butadiene. Ottawa, Ontario,Canadian Petroleum Products Institute.

Crouch CN, Pullinger DH, Gaunt IF (1979) Inhalation toxicity studieswith 1,3-butadiene — 2. 3 month toxicity study in rats. American

Industrial Hygiene Association journal , 40:796–802.

Csanády GA, Guengerich FP, Bond JA (1992) Comparison of thebiotransformation of 1,3-butadiene and its metabolite, butadienemonoepoxide, by hepatic and pulmonary tissues from humans, rats,and mice. Carcinogenesis, 13:1143–1153.

Csanády GA, Kreuzer PE, Baur C, Filser JG (1996) A physiologicaltoxicokinetic model for 1,3-butadiene in rodents and man: bloodconcentrations of 1,3-butadiene, its metabolically formed epoxides, andof haemoglobin adducts — relevance of glutathione depletion.Toxicology, 113:300–305.

Cunningham MJ, Choy WN, Arce GT, Rickard LB, Vlachos DA, KinneyLA, Sarrif AM (1986) In vivo sister chromatid exchange andmicronucleus induction studies with 1,3-butadiene in B6C3F 1 mice andSprague-Dawley rats. Mutagenesis, 6:449–452.

Darnell KR, Lloyd AC, Winer AM, Pitts JN Jr (1976) Reactivity scalefor atmospheric hydrocarbons based on reaction with hydroxyl radical.Environmental science and technology, 10:692–696.

Delzell E, Sathiakumar N, Macaluso M, Hovinga M, Larson R, Barone F,Beall C, Cole P, Julian J, Muir DCF (1995) A follow-up study of

synthetic rubber workers. Prepared for the International Institute ofSynthetic Rubber Workers, 2 October 1995.

Delzell E, Macaluso M, Lally C, Cole P (1996) Mortality study of

synthetic rubber workers: additional analyses of data on monomer peaks

and employment in certain work areas. Prepared for the InternationalInstitute of Synthetic Rubber Workers, 16 October 1996.

de Meester C, Poncelet F, Roberfroid M, Mercier M (1978) Mutagenicityof butadiene and butadiene monoxide. Biochemical and biophysical

research communications, 80:298 [cited in de Meester et al., 1980].

de Meester C, Poncelet F, Roberfroid M, Mercier M (1980) Themutagenicity of butadiene towards Salmonella typhimurium. Toxicology

letters, 6:125–130.

Deutschmann S, Laib RJ (1989) Concentration-dependent depletion ofnon-protein sulfhydryl (NPSH) content in lung, heart and liver tissue ofrats and mice after acute inhalation exposure to butadiene. Toxicology

letters, 45:175–183.

Divine BJ, Hartman CM (1996) Mortality update of butadiene productionworkers. Toxicology, 113:169–181.

Divine BJ, Wendt JK, Hartman CM (1993) Cancer mortality amongworkers at a butadiene production facility. In: Sorsa M, Peltonen K,Vainio H, Hemminki K, eds. Butadiene and styrene: Assessment of



43

Doerr JK, Hooser SB, Smith BJ, Sipes IG (1995) Ovarian toxicity of 4-vinylcyclohexene and related olefins in B6C3F 1 mice: role ofdiepoxides. Chemical research in toxicology, 8:963–969.

Doerr JK, Hollis EA, Sipes IG (1996) Species difference in the ovariantoxicity of 1,3-butadiene epoxides in B6C3F 1 mice and Sprague-Dawleyrats. Toxicology, 113:128–136.

Downs T, Pier S, Crane M, Yim K, Kim K (1993) Cause-specificmortality in a cohort of 1000 ABS workers. In: International Symposium

on Health Hazards of Butadiene and Styrene — Abstracts. 18–21 April1993, Espoo, Finland, p. 72.

Duescher RJ, Elfarra AA (1992) 1,3-Butadiene oxidation by humanmyeloperoxidase. Role of chloride ion in catalysis of divergentpathways. Journal of biological chemistry, 267(28):19 859–19 865.

Duescher RJ, Elfarra AA (1994) Human liver microsomes are efficientcatalysts of 1,3-butadiene oxidation: evidence for major roles bycytochromes P450 2A6 and 2E1. Archives of biochemistry and

biophysics, 311(2):342–349.

Duffy BL, Nelson PF (1996) Non-methane exhaust composition in theSydney Harbour tunnel: a focus on benzene and 1,3-butadiene.Atmospheric environment, 30(15):2759–2768.

Eisenreich SJ, Looney BB, Thornton JD (1981) Airborne organiccontaminants in the Great Lakes ecosystem. Environmental science and

technology, 15:30–38.

Elfarra AA, Krause RJ, Selzer RR (1996) Biochemistry of 1,3-butadienemetabolism and its relevance to 1,3-butadiene-induced carcinogenicity.Toxicology, 113:23–30.

Environment Canada (1994) Underground garage air quality assessment

program. Prepared by L. Graham, D. Rosenblatt, and P. Barton,Technology Development Directorate, Environment Canada, for S.Lamy, Environmental Health Directorate, Health Canada, Ottawa,Ontario (MSED Report No. 94-29).

Environment Canada (1996a) Summary report — 1994. Hull, Quebec,Environment Canada, Pollution Data Branch, National Pollutant ReleaseInventory (NPRI).

Environment Canada (1996b) National Analysis of Trends in

Emergencies System (NATES) database. Hull, Quebec, EnvironmentCanada, Environmental Emergencies Branch.

Environment Canada (1997) Summary report — 1995. Hull, Quebec,Environment Canada, Pollution Data Branch, National Pollutant ReleaseInventory (NPRI).

Environment Canada (1998) Priority Substances List supporting

document (environmental assessment) — 1,3-butadiene. Hull, Quebec,Environment Canada, Commercial Chemicals Evaluation Branch.

Epicure (1993) Epicure: Risk regression and data analysis software.

Seattle, WA, HiroSoft International Corporation.

European Chemicals Bureau (2001) European Union risk assessment:

1,3-Butadiene. Prepared by Institute for Health and ConsumerProtection, European Chemicals Bureau. Luxembourg, Office forOfficial Publications of the European Communities (to be published).

Gerin M, Siemiatycki J (1998) An evaluation of the exposure

assessment methods developed at the University of Alabama for the

study of cancer among synthetic rubber workers. Prepared for PrioritySubstances Program, Health Canada, Ottawa, Ontario.

Government of Canada (2000) Canadian Environmental Protection Act,

1999. Priority Substances List Assessment Report. 1,3-Butadiene.

Ottawa, Ontario, Environment Canada and Health Canada.

Hackett PL, Sikov MR, Mast TJ, Brown MG, Buschbom RL, Clark ML,Decker JR, Evanoff JJ, Rommereim RL, Rowe SE, Westerberg RB(1987a) Inhalation developmental toxicology studies of 1,3-butadiene in

the rat. Richland, WA, Pacific Northwest Laboratory (Report No. PNL-6414/UC-48) [cited in Morrissey et al., 1990].

Hackett PL, Sikov MR, Mast TJ, Brown MG, Buschbom RL, Clark ML,Decker JR, Evanoff JJ, Rommereim RL, Rowe SE, Westerberg RB(1987b) Inhalation developmental toxicology studies: teratology study of

1,3-butadiene in mice. Richland, WA, Pacific Northwest Laboratory(Report No. PNL-6412/UC-48) [cited in Morrissey et al., 1990].

Hallberg LM, Bechtold WE, Grady J, Legator MS, Au WW (1997)Abnormal DNA repair activities in lymphocytes of workers exposed to1,3-butadiene. Mutation research, 383:213–221.

Hamilton-Wentworth (1997) Human health risk assessment for priority air

pollutants. Regional Municipality of Hamilton-Wentworth, Ontario,Hamilton-Wentworth Air Quality Initiative, December.

Hayes RB, Xi L, Bechtold WE, Rothman N, Yao M, Henderson R,Zhang L, Smith MT, Zhang D, Wiemels J, Dosemeci M, Yin S, O’NeillJP (1996) hprt mutation frequency among workers exposed to 1,3-butadiene in China. Toxicology, 113:100–105.

Hazleton Laboratories Europe Ltd. (1981a) The toxicity and

carcinogenicity of butadiene gas administered to rats by inhalation for

approximately 24 months. Final report. Prepared by P.E. Owen.Harrogate (Report No. 2653-522/2).

Hazleton Laboratories Europe Ltd. (1981b) 1,3-Butadiene: Inhalation

teratogenicity study in the rat. Final report. Prepared by L.F.H. Irvine.Harrogate (Report No. 2788-522/3).

Hazleton Laboratories Europe Ltd. (1982) 1,3-Butadiene: Inhalation

teratogenicity study in the rat. Addendum to final report. Prepared byL.F.H. Irvine. Harrogate (Report No. 2788-522/3).

Henderson RF, Bechtold WE, Sabourin PJ, Maples KR, Dahl AR (1993)Species differences in the metabolism of 1,3-butadiene in vivo. In:Sorsa M, Peltonen K, Vainio H, Hemminki K, eds. Butadiene and

styrene: Assessment of health hazards. Lyon, International Agency forResearch on Cancer, pp. 57–64 (IARC Scientific Publications No. 127).

Henderson RF, Thornton-Manning JR, Bechtold WE, Dahl AR (1996)Metabolism of 1,3-butadiene: species differences. Toxicology,113:17–22.

Himmelstein MW, Turner MJ, Asgharian B, Bond JA (1994) Comparisonof blood concentrations of 1,3-butadiene and butadiene epoxides inmice and rats exposed to 1,3-butadiene by inhalation. Carcinogenesis,15(8):1479–1486.

Himmelstein MW, Asgharian B, Bond JA (1995) High concentrations ofbutadiene epoxides in livers and lungs of mice compared to ratsexposed to 1,3-butadiene. Toxicology and applied pharmacology,132:281–288.

Himmelstein MW, Acquavella JF, Recio L, Medinsky MA, Bond JA(1997) Toxicology and epidemiology of 1,3-butadiene. Critical reviews in

toxicology, 27(1):1–108.

Howard PH, Boethling RS, Jarvis WF, Meylan WM, Michalenko EM(1991) Handbook of environmental degradation rates. Chelsea, MI, LewisPublishers Inc.

Howe RB (1995a) THC: A computer program to compute a reference dose

from continuous animal toxicity data using the benchmark dose method.Ruston, LA, ICF Kaiser Engineers, Inc.

Howe RB (1995b) THRESH: A computer program to compute a reference

dose from quantal animal toxicity data using the benchmark dose method.Ruston, LA, ICF Kaiser Engineers, Inc.

Howe RB, Crump KS (1982) Global82: A computer program to

extrapolate quantal animal toxicity data to low doses. Ruston, LA,Science Research Systems.

HSE (1992) Methods for the determination of hazardous substances

(MDHS) 53 — Pumped, molecular sieve. United Kingdom Health andSafety Executive. London, Her Majesty’s Stationery Office [cited inIARC, 1999].


44

IARC (1992) Occupational exposures to mists and vapours from strong

inorganic acids; and other industrial chemicals. Lyon, InternationalAgency for Research on Cancer (IARC Monographs on the Evaluationof Carcinogenic Risks to Humans, Vol. 54).

IARC (1999) Re-evaluation of some organic chemicals, hydrazine and

hydrogen peroxide (Part one) . Lyon, International Agency for Researchon Cancer (IARC Monographs on the Evaluation of Carcinogenic Risksto Humans, Vol. 71).

IPCS (1993) International Chemical Safety Card — 1,3-Butadiene.

Geneva, World Health Organization, International Programme onChemical Safety (ICSC 0017).

Irons RD, Pyatt DW (1998) Dithiocarbamates as potential confoundersin butadiene epidemiology. Carcinogenesis, 49(4):539–542.

Irons RD, Smith CN, Stillman WS, Shah RS, Steinhagen WH,Leiderman LJ (1986a) Macrocytic-megaloblastic anemia in male B6C3F 1

mice following chronic exposure to 1,3-butadiene. Toxicology and

applied pharmacology, 83:95–100.

Irons RD, Smith CN, Stillman WS, Shah RS, Steinhagen WH,Leiderman LJ (1986b) Macrocytic-megaloblastic anemia in male NIHSwiss mice following repeated exposure to 1,3-butadiene. Toxicology

and applied pharmacology, 85:450–455.

Irons RD, Oshimura M, Barrett JC (1987) Chromosome aberrations inmouse bone marrow cells following in vivo exposure to 1,3-butadiene.Carcinogenesis, 8:1711–1714.

Irons RD, Cathro HP, Stillman WS, Steinhagen WH, Shah RS (1989)Susceptibility to 1,3-butadiene-induced leukemogenesis correlates withendogenous ecotropic retroviral background in the mouse. Toxicology

and applied pharmacology, 101:170–176.

Irons RD, Le AT, Som DB, Stillman WS (1995) 2'3'-Dideoxycytidine-induced thymic lymphoma correlates with species-specific suppressionof a subpopulation of primitive hematopoietic progenitor cells in mousebut not rat or human bone marrow. Journal of clinical investigation,95:2777–2782.

Irons RD, Colagiovanni DB, Stillman WS (1996) Murine thymiclymphoma is associated with a species-specific hematopoieticprogenitor cell subpopulation. Toxicology, 113:59–67.

Jauhar PP, Henika PR, MacGregor JT, Wehr CM, Shelby MD, MurphySA, Margolin BH (1988) 1,3-Butadiene: induction of micronucleatederythrocytes in the peripheral blood of B6C3F 1 mice exposed byinhalation for 13 weeks. Mutation research, 209:171–176.

Jelitto B, Vangala RR, Laib RJ (1989) Species differences in DNAdamage by butadiene: Role of diepoxybutane. Archives of toxicology

supplement, 13:246–249.

Kelsey KT, Wiencke JK, Ward J, Bechtold W, Fajen J (1995) Sister-chromatid exchanges, glutathione S-transferase 2 deletion andcytogenetic sensitivity to diepoxybutane in lymphocytes frombutadiene monomer production workers. Mutation research,335:267–273.

Kenaga EE (1980) Predicted bioconcentration factors and soil sorptioncoefficients of pesticides and other chemicals. Ecotoxicology and

environmental safety, 4:26–38.

Klöpffer W, Haag F, Kohl E-G, Frank R (1988) Testing of the abioticdegradation of chemicals in the atmosphere: The smog chamberapproach. Toxicology and environmental safety, 15:298–319.

Koivisto P, Sorsa M, Pacchierotti F, Peltonen K (1997) 32P-postlabelling/HPLC assay reveals an enantioselective adduct formationin N7 guanine residues in vivo after 1,3-butadiene inhalation exposure.Carcinogenesis, 18(2):439–443.

Koivisto P, Adler I-D, Pacchierotti F, Peltonen K (1998) DNA adducts inmouse testes and lung after inhalation exposure to 1,3-butadiene.Mutation research, 397:3–10.

Krause RJ, Elfarra AA (1997) Oxidation of butadiene monoxide to meso-and (±)-diepoxybutane by cDNA-expressed human cytochrome P450sand by mouse, rat, and human liver microsomes: evidence forpreferential hydration of meso-diepoxybutane in rat and human livermicrosomes. Archives of biochemistry and biophysics, 337(2):176–184.

Krause RJ, Sharer JE, Elfarra AA (1997) Epoxide hydrolase-dependentmetabolism of butadiene monoxide to 3-butene-1,2-diol in mouse, rat,and human liver. Drug metabolism and disposition, 25(8):1013–1015.

Kreiling R, Laib RJ, Bolt HM (1986) Alkylation of nuclear proteins andDNA after exposure of rats and mice to [1,4- 14C]1,3-butadiene.Toxicology letters, 30:131–136.

Kreuzer PE, Kessler W, Welter HF, Baur C, Filser JG (1991) Enzymespecific kinetics of 1,2-epoxybutene-3 in microsomes and cytosol fromlivers of mouse, rat, and man. Archives of toxicology, 65:59–67.

Labstat, Inc. (1995) An assessment of the chemical toxicity of the smoke

from Canadian cigarettes: Method development and analytical results for

ammonia, pyridine, 1-3 butadienne [sic] and vinyl chloride. Final report.

30 November 1995. Determined under contract with Health Canada(H1021-4-9127/02-SS).

Lähdetie J, Grawé J (1997) Flow cytometric analysis of micronucleusinduction in rat bone marrow polychromatic erythrocytes by 1,2:3,4-diepoxybutane, 3,4-epoxy-1-butene, and 1,2-epoxybutane-3,4-diol.Cytometry, 28:228–235.

Lähdetie J, Peltonen K, Sjoblöm T (1997) Germ cell mutagenicity ofthree metabolites of 1,3-butadiene in the rat: induction of spermatidmicronuclei by butadiene mono-, di-, and diolepoxides in vivo.Environmental and molecular mutagenesis, 29:230–239.

Landi S, Ponzanelli I, Hirvonen A, Norppa H, Barale R (1996) Repeatedanalysis of sister chromatid exchange induction by diepoxybutane incultured human lymphocytes: effect of glutathione S-transferase T1and M1 genotype. Mutation research, 351:79–85.

Larionov LT, Shtessel TA, Nuselman EN (1934) Concerning the effectof butadiene, pseudobutylene and isoprene. Kazanskij Meditsinskij

Zhurnal , 5:440–445.

Leavens TL, Farris GM, James RA, Shah R, Wong VA, Marshall MW,Bond JA (1997) Genotoxicity and cytotoxicity in male B6C3F 1 micefollowing exposure to mixtures of 1,3-butadiene and styrene.Environmental and molecular mutagenesis, 29:335–345.

Legator MS, Au WW, Ammenheuser M, Ward JB Jr (1993) Elevatedsomatic cell mutant frequencies and altered DNA repair responses innonsmoking workers exposed to 1,3-butadiene. In: Sorsa M, Peltonen K,Vainio H, Hemminki K, eds. Butadiene and styrene: Assessment of


Leiderman LJ, Stillman WS, Shah RS, Steinhagen WH, Irons RD (1986)Altered hematopoietic stem cell development in male B6C3F 1 micefollowing exposure to 1,3-butadiene. Experimental and molecular

pathology, 44:50–56.

Ligocki MP, Fieber JL, Ball JC, Pezda SA, Heuss JM, Paul RT, WimetteHJ (1994) Sources, projected emission trends and exposure issues for

1,3-butadiene. Presentation at the 87th Annual Meeting and Exhibition ofthe Air and Waste Management Association, Cincinnati, OH, June.

Linet MS, Stewart W, Van Natta ML, McCaffrey LD, Szklo M (1987)Comparison of methods for determining occupational exposure in acase–control interview study of chronic lymphocytic leukemia. Journal

of occupational medicine, 29(2):136–141.

Lyman WJ, Reehl WF, Rosenblatt DH (1982) Handbook of chemical

property estimation methods. New York, NY, McGraw-Hill Book Co.

Mabon N, Moorthy B, Randerath E, Randerath K (1996) Monophosphate32P-postlabeling assay of DNA adducts from 1,2:3,4-diepoxybutane, themost genotoxic metabolite of 1,3-butadiene: in vitro methodologicalstudies and in vivo dosimetry. Mutation research, 371:87–104.


45

Macaluso M, Delzell E, Sanders M, Larson R (1997) Historical

estimation of exposure to butadiene and styrene among

synthetic rubber workers. Prepared for the International Instituteof Synthetic Rubber Workers, 22 August 1997.

Mackay D (1991) Multimedia environmental models: the fugacity

approach. Chelsea, MI, Lewis Publishers Inc.

Mackay D, Paterson S (1991) Evaluating the multimedia fate oforganic chemicals: a Level III fugacity model. Environmental

science and technology, 25:427.

Mackay D, Shiu WY, Ma KC (1993) Illustrated handbook of

physical-chemical properties and environmental fate of organic

compounds. Vol. III. Chelsea, MI, Lewis Publishers Inc.

Maniglier-Poulet C, Cheng X, Ruth JA, Ross D (1995)Metabolism of 1,3-butadiene monoxide in mouse and humanbone marrow cells. Chemico-biological interactions,97:119–129.

Matanoski GM, Santos-Burgoa C, Schwartz L (1990) Mortality ofa cohort or workers in the styrene-butadiene polymermanufacturing industry (1943–1982). Environmental health

perspectives, 86:107–117.

Matanoski G, Francis M, Correa-Villaseñor A, Elliott E, Santos-Burgoa C, Schwartz L (1993) Cancer epidemiology amongstyrene-butadiene rubber workers. In: Sorsa M, Peltonen K,Vainio H, Hemminki K, eds. Butadiene and styrene: Assessment

of health hazards. Lyon, International Agency for Research onCancer, pp. 363–374 (IARC Scientific Publications No. 127).

Matanoski G, Elliott E, Tao X, Francis M, Correa-Villasenor A,Santos-Burgoa C (1997) Lymphohematopoietic cancers andbutadiene and styrene exposure in synthetic rubber manufacture.Annals of the New York Academy of Science, 837:157–169.

McKone TE, Daniels JI, Chiao FF, Hsieh DPH (1993) Intermedia

transfer factors of fifteen toxic pollutants released to air basins

in California. Livermore, CA, Lawrence Livermore NationalLaboratory (UCRL-CR-115620).

McMichael AJ, Spirtas R, Kupper LL (1974) An epidemiologicstudy of mortality within a cohort of rubber workers, 1964–72.Journal of occupational medicine, 16:458–464.

McMichael AJ, Spirtas R, Gamble JF, Tousey PM (1976)Mortality among rubber workers: Relationship to specific jobs.Journal of occupational medicine, 18:178–185.

McNeal TP, Breder CV (1987) Headspace gas chromatographicdetermination of residual 1,3-butadiene in rubber-modifiedplastics and its migration from plastic containers into selectedfoods. Journal of the Association of Official Analytical Chemists,70(1):18–21.

Meinhardt TJ, Lemen RA, Crandall MS, Young RJ (1982)Environmental epidemiologic investigation of the styrene-butadiene rubber industry. Scandinavian journal of work,

environment and health, 8:250–259.

Melnick RL, Huff JE (1992) 1,3-Butadiene: Toxicity and carcino-genicity in laboratory animals and in humans. Reviews of

environmental contamination and toxicology, 124:111–144.

Melnick RL, Huff JE, Roycroft JH, Chou BJ, Miller RA (1990)Inhalation toxicology and carcinogenicity of 1,3-butadiene inB6C3F1 mice following 65 weeks of exposure. Environmental

health perspectives, 86:27–36.

Meng Q, Recio L, Reilly AA, Wong BA, Bauer M, Walker VE(1998) Comparison of the mutagenic potency of 1,3-butadieneat hprt locus of T-lymphocytes following inhalation exposure offemale B6C3F1 mice and F344 rats. Carcinogenesis,19(6):1019–1027.

Meng Q, Henderson RF, Walker DM, Bauer MJ, Reilly AA,Walker VE (1999) Mutagenicity of the racemic mixtures ofbutadiene monoepoxide and butadiene diepoxide at the Hprt

locus of T-lymphocytes following inhalation exposures of femalemice and rats. Mutation research, 429:127–140.

Meng Q, Singh N, Heflich RH, Bauer MJ, Walker VE (2000)Comparison of the mutations at Hprt exon 3 of T-lymphocytesfrom B6C3F1 mice and F344 rats exposed by inhalation to 1,3-butadiene or the racemic mixture of 1,2:3,4-diepoxybutane.Mutation research, 464(2):169–184.

MOEE (1995) Technical memorandum — 1995 results of the

Mobile TAGA 6000: 1,3-butadiene levels in Sarnia and selected

areas in Ontario. Toronto, Ontario, Ontario Ministry ofEnvironment and Energy.

Morrissey RE, Schwetz BA, Hackett PL, Sikov MR, Hardin BD,McClanahan BJ, Decker JR, Mast TJ (1990) Overview of repro-ductive and developmental toxicity studies of 1,3-butadiene inrodents. Environmental health perspectives, 86:79–84.

Nauhaus SK, Fennell TR, Asgharian B, Bond JA, Sumner SCJ(1996) Characterization of urinary metabolites from Sprague-Dawley rats and B6C3F1 mice exposed to [1,2,3,4-13C]butadiene.Chemical research in toxicology, 9:764–773.

Nelson HH, Wiencke JK, Christiani DC, Cheng TJ, Zuo Z-F,Schwartz BS, Lee B-K, Spitz MR, Wang M, Xu X, Kelsey KT(1995) Ethnic differences in the prevalence of the homozygousdeleted genotype of glutathione S-transferase theta.Carcinogenesis, 16(5):1243–1245.

Neumann H-G, Albrecht O, Van Dorp C, Zwirner-Baier I (1995)Macromolecular adducts caused by environmental chemicals.Clinical chemistry, 41(12):1835–1840.

Nikiforova AA, Ripp GK, Taskayev II (1969) Action of 1,3-butadiene on the structural elements of kidneys and heart.Nauchnye Trudy, Omskii Meditsinskii Institut, 88:166–169.

NIOSH (1994) NIOSH manual of analytical methods (NMAM), 4thed. Cincinnati, OH, US Department of Health and HumanServices, Public Health Service, Centers for Disease Control andPrevention, National Institute for Occupational Safety andHealth.

Norppa H, Hirvonen A, Jarventaus H, Uuskula M, Tasa G,Ojajarvi A, Sorsa M (1995) Role of GSTT1 and GSTM1

genotypes in determining individual sensitivity to sisterchromatid exchange induction by diepoxybutane in culturedhuman lymphocytes. Carcinogenesis, 16(6):1261–1264.

NTP (1984) NTP technical report on the toxicology and carcino-

genesis studies of 1,3-butadiene (CAS No. 106-99-0) in B6C3F1

mice (inhalation studies). Research Triangle Park, NC, USDepartment of Health and Human Services, National ToxicologyProgram (Technical Report No. 288).

NTP (1993) NTP technical report on the toxicology and carcino-

genesis studies of 1,3-butadiene (CAS No. 106-99-0) in B6C3F1

mice (inhalation studies). Research Triangle Park, NC, USDepartment of Health and Human Services, National ToxicologyProgram (Technical Report No. 434).


46

OECD (1996) Draft risk assessment of butadiene. Paris,Organisation for Economic Co-operation and Development,March.

Osterman-Golkar SM, Bond JA, Ward JB, Legator MS (1993) Useof haemoglobin adducts for biomonitoring exposure to 1,3-butadiene. In: Sorsa M, Peltonen K, Vainio H, Hemminki K, eds.Butadiene and styrene: Assessment of health hazards. Lyon,International Agency for Research on Cancer, pp. 185–193(IARC Scientific Publications No. 127).

Osterman-Golkar S, Peltonen K, Anttinen-Klemetti T, HindsøLandin H, Zorcec V, Sorsa M (1996) Haemoglobin adducts asbiomarkers of occupational exposure to 1,3-butadiene.Mutagenesis, 11(2):145–149.

Owen PE, Glaister JR (1990) Inhalation toxicity andcarcinogenicity of 1,3-butadiene in Sprague-Dawley rats.Environmental health perspectives, 86:19–25.

Owen PE, Glaister JR, Gaunt IF, Pullinger DH (1987) Inhalationtoxicity studies with 1,3-butadiene: 3. Two year toxicity/carcinogenicity study in rats. American Industrial Hygiene

Association journal, 48:407–413.

Pacchierotti F, Tiveron C, Ranaldi R, Bassani B, Cordelli E,Leter G, Spanò M (1998a) Reproductive toxicity of 1,3-butadiene in the mouse: cytogenetic analysis of chromosomeaberrations in first-cleavage embryos and flow cytometricevaluation of spermatogonial cell killing. Mutation research,397:55–66.

Pacchierotti F, Adler I-D, Anderson D, Brinkworth M, DemopoulosNA, Lähdetie J, Osterman-Golkar S, Peltonen K, Russo A, TatesA, Waters R (1998b) Genetic effects of 1,3-butadiene andassociated risk for heritable damage. Mutation research,397:93–115.

Pakdel H, Couture G, Roy C, Masson A, Locat J, Gelinas P,Lesage S (1992) Developing methods for the analysis of toxicchemicals in soil and groundwater: the case of Ville Mercier,Quebec, Canada. In: Lesage S, Jackson R, eds. Groundwater

contamination and analysis at hazardous waste sites. New York,NY, Marcel Dekker, Inc., pp. 381–421 (Environmental Scienceand Pollution Control Series).

Pelin K, Hirvonen A, Norppa H (1996) Influence of erythrocyteglutathione S-transferase T1 on sister chromatid exchangesinduced by diepoxybutane in cultured human lymphocytes.Mutagenesis, 11(2):213–215.

Pellizzari ED, Michael LC, Thomas KW, Shields PG, Harris C(1995) Identification of 1,3-butadiene, benzene, and othervolatile organics from wok oil emissions. Journal of exposure

analysis and environmental epidemiology, 5(1):77–87.

Peltonen K, Koivisto P, Neagu I, Kostiainen R, Kilpelainen I,Sorsa M (1993) Estimating internal dose of 1,3-butadiene:preliminary data on use of modified purine bases as markers ofexposure. In: Sorsa M, Peltonen K, Vainio H, Hemminki K, eds.Butadiene and styrene: Assessment of health hazards. Lyon,International Agency for Research on Cancer, pp. 119–126(IARC Scientific Publications No. 127).

Pérez HL, Lähdetie J, Landin HL, Kilpelainen I, Koivisto P,Peltonen K, Osterman-Golkar S (1997) Haemoglobin adducts ofepoxybutanediol from exposure to 1,3-butadiene or butadieneepoxides. Chemico-biological interactions, 105:181–198.

Peto R, Pike MC, Berstein L, Gold LS, Ames BN (1984) TheTD50: A proposed general convention for the numericaldescription of the carcinogenic potency of chemicals in chronic-

exposure animal experiments. Environmental health

perspectives, 58:1–8.

Portier CJ, Bailer AJ (1989) Testing for increased carcinogenicityusing a survival-adjusted quantal response test. Fundamental

and applied toxicology, 12:731–737.

Preston DL, Kato H, Kopecky KJ, Fujita S (1987) Studies of themortality of a-bomb survivors: 8. Cancer mortality, 1950–1982.Radiation research, 111:151–178.

Przygoda RT, Bird MG, Whitman FT, Wojcik NC, McKee RH(1993) Induction of micronuclei in mice and hamsters by 1,3-butadiene. Environmental and molecular mutagenesis,21(Suppl. 22):56.

Recio L, Meyer KG (1995) Increased frequency of mutations atA:T base pairs in the bone marrow of B6C3F1 lacI transgenicmice exposed to 1,3-butadiene. Environmental and molecular

mutagenesis, 26:1–8.

Recio L, Osterman-Golkar S, Csanády GA, Turner MJ, Myhr B,Moss O, Bond JA (1992) Determination of mutagenicity intissues of transgenic mice following exposure to 1,3-butadieneand N-ethyl-N-nitrosourea. Toxicology and applied

pharmacology, 117:58–64.

Recio L, Bond JA, Pluta LJ, Sisk SC (1993) Use of transgenicmice for assessing the mutagenicity of 1,3-butadiene in vivo. In:Sorsa M, Peltonen K, Vainio H, Hemminki K, eds. Butadiene

and styrene: Assessment of health hazards. Lyon, InternationalAgency for Research on Cancer, pp. 235–243 (IARC ScientificPublications No. 127).

Recio L, Meyer KG, Pluta LJ, Moss OR, Saranko CJ (1996)Assessment of 1,3-butadiene mutagenicity in the bone marrowof B6C3F1 lacI transgenic mice (Big Blue®): a review ofmutational spectrum and lacI mutant frequency after a 5-day625 ppm 1,3-butadiene exposure. Environmental and molecular


Recio L, Henderson RF, Pluta LJ, Meyer K, Saranko CJ, SteenA-M (1998) Analysis of mutagenicity and determination ofmutational spectrum in rodent and human cells to assess theroles of epoxybutene and diepoxybutane in mediating the invivo genotoxicity of 1,3-butadiene. In: Proceedings of the 14th

Health Effects Institute Annual Conference. Cambridge, MA,Health Effects Institute, p. 49.

Ripp GK (1967) Hygiene basis for the permissible concentrationof butadiene in the atmosphere. Biologicheskoe Deistvie i

Gigienicheskoe Znachenie Atmosfernykh Zagryaznenii,10:33–54.

Ristau C, Deutschmann S, Laib RJ, Ottenwalder H (1990)Detection of diepoxybutane-induced DNA–DNA crosslinks bycesium trifluoracetate (CsTFA) density-gradient centrifugation.Archives of toxicology, 64:343–344.

Ross D, Siegel D, Schattenberg DG, Sun XM, Moran JL (1996)Cell-specific activation and detoxification of benzenemetabolites in mouse and human bone marrow: identification oftarget cells and a potential role for modulation of apoptosis inbenzene toxicity. Environmental health perspectives, 104(Suppl.6):1177–1182.

Russo A, Nogara C, Renzi L, Tommasi AM (1997) Micronucleusinduction in germ and somatic cells of the mouse after exposureto the butadiene metabolites diepoxybutane and epoxybutene.Mutation research, 390:129–139.


47

Sabourin PJ, Burka LT, Bechtold WE, Dahl AR, Hoover MD,Chang IY, Henderson RF (1992) Species differences in urinarybutadiene metabolites; identification of 1,2-dihydroxy-4-(N-acetylcysteinyl)butane, a novel metabolite of butadiene.Carcinogenesis, 13(9):1633–1638.

Santos-Burgoa C, Matanoski GM, Zeger S, Schwartz L (1992)Lymphohematopoietic cancer in styrene-butadienepolymerization workers. American journal of epidemiology,136:843–854.

Saranko CJ, Recio L (1998) The butadiene metabolite, 1,2:3,4-diepoxybutane, induces micronuclei but is only weaklymutagenic at lacI in the Big Blue® Rat2 lacI transgenic cell line.Environmental and molecular mutagenesis, 31:32–40.

Saranko CJ, Pluta LJ, Recio L, Henderson RF (1998) In vivo andin vitro mutagenicity spectrum of the butadiene metabolite 1,2-epoxybutene. Proceedings of the American Association of

Cancer Research, 39:330 (abstract).

Sasiadek M, Järventaus H, Sorsa M (1991a) Sister-chromatidexchanges induced by 1,3-butadiene and its epoxides in CHOcells. Mutation research, 263:47–50.

Sasiadek M, Norppa H, Sorsa M (1991b) 1,3-Butadiene and itsepoxides induce sister-chromatid exchanges in humanlymphocytes in vitro . Mutation research, 261:117–121.

Sathiakumar N, Delzell E, Hovinga M, Macaluso M, Julian JA,Larson R, Cole P, Muir DCF (1998) Mortality from cancer andother causes of death among synthetic rubber workers.Occupational and environmental medicine, 55:230–235.

Schattenberg DG, Stillman WS, Gruntmeir JJ, Helm KM, IronsRD, Ross D (1994) Peroxidase activity in murine and humanhematopoietic progenitor cells: potential relevance to benzene-induced toxicity. Molecular pharmacology, 46:346–351.

Schuetzle D, Siegl WO, Jensen TE, Dearth MA, Kaiser EW,Gorse R, Kreucher W, Kulik E (1994) The relationship betweengasoline composition and vehicle hydrocarbon emissions: areview of current studies and future research needs.Environmental health perspectives, 102(Suppl. 4):3–12.

Sernau R, Cavagnaro J, Kehn P (1986) 1,3-Butadiene as an S9activation-dependent gaseous positive control substance inL5178Y cell mutation assays. Environmental mutagenesis,8(Suppl.):75 (Abstract 203).

Sharer JE, Duescher RJ, Elfarra AA (1992) Species and tissuedifferences in the microsomal oxidation of 1,3-butadiene andthe glutathione conjugation of butadiene monoxide in mice andrats. Drug metabolism and disposition, 20:658–664.

Sharief Y, Brown AM, Backer LC, Campbell JA, Westbrook-Collins B, Stead AG, Allen JW (1986) Sister chromatid exchangeand chromosome aberration analyses in mice after in vivo

exposure to acrylonitrile, styrene, or butadiene monoxide.Environmental mutagenesis, 8:439–448.

Shelby MD (1990) Results of NTP-sponsored mouse cytogeneticstudies on 1,3-butadiene, isoprene, and chloroprene.Environmental health perspectives, 86:71–73.

Shields PG, Xu GX, Blot WJ, Fraumeni JF Jr, Trivers GE,Pellizzari ED, Qu YH, Gao YT, Harris CC (1995) Mutagens fromheated Chinese and U.S. cooking oils. Journal of the National

Cancer Institute, 887(110):836–841.

Shugaev B (1969) Concentrations of hydrocarbons in tissues as ameasure of toxicity. Archives of environmental health,18:878–882.

Siemiatycki J (1991) Risk factors for cancer in the workplace.Boca Raton, FL, CRC Press.

Sisk SC, Pluta LJ, Bond JA, Recio L (1994) Molecular analysis oflacI mutants from bone marrow of B6C3F1 transgenic micefollowing inhalation exposure to 1,3-butadiene. Carcinogenesis,15:471–477.

Sjöblom T, Lähdetie J (1996) Micronuclei are induced in ratspermatids in vitro by 1,2,3,4-diepoxybutane but not 1,2-epoxy-3-butene and 1,2-dihydroxy-3,4-epoxybutane. Mutagenesis,11(5):525–528.

Sorsa M, Autio J, Demopoulos NA, Jarventaus H, Rossner P,Sram RJ, Stephanou G, Vlachodimitropoulos D (1994) Humancytogenetic biomonitoring of occupational exposure to 1,3-butadiene. Mutation research, 309:321–326.

Sorsa M, Osterman-Golkar S, Peltonen K, Saarikoski ST, Šram R(1996a) Assessment of exposure to butadiene in the processindustry. Toxicology, 113:77–83.

Sorsa M, Peltonen K, Anderson D, Demopoulos NA, NeumannH-G, Osterman-Golkar S (1996b) Assessment of environmentaland occupational exposures to butadiene as a model for riskestimation of petrochemical emissions. Mutagenesis,11(1):9–17.

Spano M, Bartoleschi C, Cordelli E, Leter G, Segre L,Mantovani A, Fazzi P, Pacchierotti F (1996) Flow cytometric andhistological assessment of 1,2:3,4-diepoxybutane toxicity onmouse spermatogenesis. Journal of toxicology and

environmental health, 47:423–441.

Šrám RJ, Rössner P, Peltonen K, Podrazilová K, Mracková G,Demopoulos NA, Stephanou G, Vladimiropoulos D, van Dam FJ,Tates AD (1998) Chromosomal aberrations, sister-chromatidexchanges, cells with high frequency of SCE, micronuclei andcomet assay parameters in 1,3-butadiene exposed workers.Mutation research, 419(1–3):145–154.

Startin JR, Gilbert J (1984) Single ion monitoring of butadienein plastics and foods by coupled mass spectrometry–automaticheadspace gas chromatography. Journal of chromatography,294:427–430.

Stephanou G, Andrianopoulos C, Vlastos D, Demopoulos NA,Russo A (1997) Induction of micronuclei and sister chromatidexchange in mouse splenocytes after exposure to the butadienemetabolite 3,4-epoxy-1-butene. Mutagenesis, 12(6):425–429.

Stephanou G, Russo A, Vlastos D, Andrianopoulos C,Demopoulos NA (1998) Micronucleus induction in somatic cellsof mice as evaluated after 1,3-butadiene inhalation. Mutation

research, 397:11–20.

Swann RL, Laskowski DA, McCall PJ, Vander Kuy K, DishburgerHJ (1983) A rapid method for the estimation of theenvironmental parameters octanol/water partition coefficient,soil sorption constant, water to air ratio and water solubility.Residue reviews, 85:17–28.

Sweeney LM, Schlosser PM, Medinsky MA, Bond JA (1997)Physiologically based pharmacokinetic modeling of 1,3-butadiene, 1,2-epoxy-3-butene, and 1,2:3,4-diepoxybutanetoxicokinetics in mice and rats. Carcinogenesis, 18(4):611–625.


48

Tates AD, van Dam FJ, de Zwart FA, van Teylingen CMM,Natarajan AT (1994) Development of a cloning assay with highcloning efficiency to detect induction of 6-thioguanine-resistantlymphocytes in spleen of adult mice following in vivo inhalationexposure to 1,3-butadiene. Mutation research, 309:299–306.

Tates AD, van Dam FJ, de Zwart FA, Darroudi F, Natarajan AT,Rössner P, Peterková K, Peltonen K, Demopoulos NA,Stephanou G, Vlachodimitropoulos D, Šrám RJ (1996)Biological effect monitoring in industrial workers from the CzechRepublic exposed to low levels of butadiene. Toxicology,113:91–99.

Tates AD, van Dam FJ, van Teylingen CMM, de Zwart FA,Zwinderman AH (1998) Comparison of induction of hprt

mutations by 1,3-butadiene and/or its metabolites 1,2-epoxybutene and 1,2,3,4-diepoxybutane in lymphocyes fromspleen of adult male mice and rats in vivo. Mutation research,397:21–36.

Thier R, Persmark M, Pemble SE, Taylor JB, Ketterer B,Guengerich FP (1994) Mutagenicity of 1,2,3,4-butadienediepoxide is altered by mammalian 2-class glutathione S-transferase. Archives of pharmacology, 349(Suppl.):R121.

Thornton-Manning JR, Dahl AR, Bechtold WE, Griffith WC,Henderson RF (1995a) Disposition of butadiene monoepoxideand butadiene diepoxide in various tissues of rats and micefollowing a low-level inhalation exposure to 1,3-butadiene.Carcinogenesis, 16(8):1723–1731.

Thornton-Manning JR, Dahl AR, Bechtold WE, Griffith WC, PeiL, Henderson RF (1995b) Gender differences in the metabolismof 1,3-butadiene in Sprague-Dawley rats following a low levelinhalation exposure. Carcinogenesis, 16(11):2875–2878.

Thornton-Manning JR, Dahl AR, Bechtold WE, Griffith WC Jr,Henderson RF (1997) Comparison of the disposition of butadieneepoxides in Sprague-Dawley rats and B6C3F1 mice following asingle and repeated exposures to 1,3-butadiene via inhalation.Toxicology, 123:125–134.

Thornton-Manning JR, Dahl AR, Allen ML, Bechtold WE, GriffithWC Jr, Henderson RF (1998) Disposition of butadiene epoxidesin Sprague-Dawley rats following exposures to 8000 ppm 1,3-butadiene: comparisons with tissue epoxide concentrationsfollowing low level exposures. Toxicological sciences,41:167–173.

Thurmond LM, Lauer LD, House RV, Stillman WS, Irons RD,Steinhagen WH, Dean JH (1986) Effect of short-term inhalationexposure to 1,3-butadiene on murine immune functions.Toxicology and applied pharmacology, 86:170–179.

Tice RR, Boucher R, Luke CA, Shelby MD (1987) Comparativecytogenetic analysis of bone marrow damage induced by maleB6C3F1 mice by multiple exposures to gaseous 1,3-butadiene.Environmental mutagenesis, 9:235–250.

Tiveron C, Ranaldi R, Bassani B, Pacchierotti F (1997) Inductionand transmission of chromosome aberrations in mouse oocytesafter treatment with butadiene epoxide. Environmental and

molecular mutagenesis, 30:403–409.

Tommasi AM, De Conti S, Dobrzynska MM, Russo A (1998)Evaluation and characterization of micronuclei in earlyspermatids of mice exposed to 1,3-butadiene. Mutation

research, 397:45–54.

Tretyakova NY, Lin Y-P, Upton PB, Sangaiah R, Swenberg JA(1996) Macromolecular adducts of butadiene. Toxicology,113:70–76.

Tretyakova NY, Chiang S-Y, Walker VE, Swenberg JA (1998a)Quantitative analysis of 1,3-butadiene-induced DNA adducts invivo and in vitro using liquid chromatography electrosprayionization tandem mass spectrometry. Journal of mass

spectrometry, 33:363–376.

Tretyakova NY, Walker VE, Swenberg JA (1998b) Formation andpersistence of DNA adducts in liver of rats and mice exposed to1,3-butadiene. In: Proceedings of the Society of Toxicology

1998 Annual Meeting. Reston, VA, Society of Toxicology, p.180 (Abstract No. 890).

US DHHS (1989) Reducing the health consequences of

smoking. 25 years of progress. A report of the Surgeon General.Rockville, MD, US Department of Health and Human Services.

US EPA (1989) Locating and estimating air emissions from

sources of 1,3-butadiene. Washington, DC, US EnvironmentalProtection Agency, Office of Air Quality Planning and Standards(EPA/450/2-89/021).

US EPA (1993) Motor vehicle-related air toxics study.Washington, DC, US Environmental Protection Agency, Office ofMobile Sources, Emission Planning and Strategies Division,April (EPA 420-R-93-005).

US FDA (1987) 1,3-Butadiene. In: Fazio T, Sherma J, eds. Food

additives analytical manual. Vol. II. A collection of analytical

methods for selected food additives. US Food and DrugAdministration. Arlington, VA, Association of Official AnalyticalChemists, pp. 56—68 [cited in IARC, 1999].

US OSHA (1990) Method 56. In: OSHA analytical methods

manual, Part 1: Organic substances. Vol. 3. Methods 55–80.Salt Lake City, UT, US Occupational Safety and HealthAdministration [cited in IARC, 1999].

Uuskula M, Jarventaus H, Hirvonen A, Sorsa M, Norppa H (1995)Influence of GSTM1 genotype on sister chromatid exchangeinduction by styrene-7,8-oxide and 1,2-epoxy-3-butene incultured human lymphocytes. Carcinogenesis, 16(4):947–950.

Van Duuren BL, Nelson N, Orris L, Palmes ED, Schmitt FL(1963) Carcinogenicity of epoxides, lactones, and peroxycompounds. Journal of the National Cancer Institute, 31:41–55[cited in IARC, 1992].

Van Duuren BL, Orris L, Nelson N (1965) Carcinogenicity ofepoxides, lactones, and peroxy compounds. Part II. Journal of

the National Cancer Institute, 35:707–717 [cited in IARC, 1992].

Van Duuren BL, Langseth L, Orris L, Teebor G, Nelson N,Kuschner M (1966) Carcinogenicity of epoxides, lactones, andperoxy compounds. IV. Tumor response in epithelial andconnective tissue in mice and rats. Journal of the National

Cancer Institute, 37:825–838.

Vangala RR, Laib RJ, Bolt HM (1993) Evaluation of DNAdamage by alkaline elution technique after inhalation exposureof rats and mice to 1,3-butadiene. Archives of toxicology,67:34–38.

Victorin K, Ståhlberg M (1988) A method for studying themutagenicity of some gaseous compounds in Salmonella

typhimurium. Environmental and molecular mutagenesis,11:65–77.

Victorin K, Busk L, Cederberg H, Magnusson J (1990) Genotoxicactivity of 1,3-butadiene and nitrogen dioxide and theirphotochemical reaction products in Drosophila and in themouse bone marrow micronucleus assay. Mutation research,228:203–209.


49

Vincent DR, Arce GT, Sarrif AM (1986) Genotoxicity of 1,3-butadiene.Assessment by the unscheduled DNA synthesis assay in B6C3F 1 miceand Sprague-Dawley rats in vivo and in vitro. Environmental

mutagenesis, 8(Suppl.):88.

Vlachodimitropoulos D, Norppa H, Autio K, Catalán J, Hirvonen A, TasaG, Uusküla M, Demopoulos NA, Sorsa M (1997) GSTT1-dependentinduction of centromere-negative and -positive micronuclei by 1,2:3,4-diepoxybutane in cultured human lymphocytes. Mutagenesis,12(5):397–403.

Walk R-A, Jenderny J, Röhrborn G, Hackenberg U (1987) Chromo-somal abnormalities and sister-chromatid exchange in bone marrow ofmice and Chinese hamsters after inhalation and intraperitonealadministration: I. Diepoxybutane. Mutation research, 182:333–342.

Walles SAS, Victorin K, Lundborg M (1995) DNA damage in lung cells invivo and in vitro by 1,3-butadiene and nitrogen dioxide and theirphotochemical reaction products. Mutation research, 328:11–19.

Ward DE, Hao WM (1992) Air toxic emissions from the burning of

biomass globally — preliminary estimates. Presented at the 85th AnnualMeeting of the Air and Waste Management Association, Kansas City,MO.

Ward EM, Fajen JM, Ruder AM, Rinsky RA, Halperin WE, Fessler-Flesch CA (1995) Mortality study of workers in 1,3-butadiene productionunits identified from a chemical workers cohort. Environmental health

perspectives, 103(6):598–603.

Ward EM, Fajen JM, Ruder AM, Rinsky RA, Halperin WE, Fessler-Flesch CA (1996) Mortality study of workers employed in 1,3-butadieneproduction units identified from a chemical workers cohort. Toxicology,113:157–168.

Ward JB Jr, Ammenheuser MM, Bechtold WE, Whorton EB Jr, LegatorMS (1994) hprt mutant lymphocyte frequencies in workers at a 1,3-butadiene production plant. Environmental health perspectives,102(Suppl. 9):79–85.

Ward JB Jr, Ammenheuser MM, Whorton EB Jr, Bechtold WE, KelseyKT, Legator MS (1996) Biological monitoring for mutagenic effects ofoccupational exposure to butadiene. Toxicology, 113:84–90.

Wiencke JK, Kelsey KT (1993) Susceptibility to induction ofchromosomal damage by metabolites of 1,3-butadiene and itsrelationship to “spontaneous” sister chromatid exchange frequencies inhuman lymphocytes. In: Sorsa M, Peltonen K, Vainio H, Hemminki K,eds. Butadiene and styrene: Assessment of health hazards. Lyon,International Agency for Research on Cancer, pp. 265–273 (IARCScientific Publications No. 127).

Wiencke JK, Pemble S, Ketterer B, Kelsey KT (1995) Gene deletion ofglutathione S-transferase 2: correlation with induced genetic damageand potential role in endogenous mutagenesis. Cancer epidemiology

biomarkers and prevention, 4:253–259.

Wilson AL, Colome SD, Tian Y (1991) Air toxics microenvironmental

exposure and monitoring study. Final report. Prepared for South CoastAir Quality Management District, El Monte, CA, and US EnvironmentalProtection Agency by Integrated Environmental Services, Irvine, CA.

Xi L, Zhang L, Wang Y, Smith MT (1997) Induction of chromosome-specific aneuploidy and micronuclei in human lymphocytes bymetabolites of 1,3-butadiene. Carcinogenesis, 18(9):1687–1693.

Xiao Y, Tates AD (1995) Clastogenic effects of 1,3-butadiene and itsmetabolites 1,2-epoxybutene and 1,2,3,4-diepoxybutane in splenocytesand germ cells of rats and mice in vivo. Environmental and molecular


APPENDIX 1 — SOURCE DOCUMENT

Government of Canada (2000)

Copies of the Canadian Environmental Protection Act

Priority Substances List Assessment Report on 1,3-butadiene(Government of Canada, 2000) are available upon request from:

Inquiry CentreEnvironment CanadaMain Floor, Place Vincent Massey351 St. Joseph Blvd.Hull, QuebecCanada K1A 0H3

or on the Internet at:

www.ec.gc.ca/cceb1/eng/public/index_e.html

Unpublished supporting documentation for the healtheffects assessment, which presents additional information, isavailable upon request from:

Environmental Health CentreRoom 104Health CanadaTunney’s PastureOttawa, OntarioCanada K1A 0L2

Sections of the Assessment Report and supportingdocumentation on genotoxicity and reproductive anddevelopmental toxicity were reviewed by D. Blakey and W.Foster, respectively, of the Environmental and OccupationalToxicology Division of Health Canada. A review of the exposureassessment included in the critical epidemiological studies wasprepared under contract by M. Gerin and J. Siemiatycki of theInstitut Armand-Frappier, University of Quebec.

In the first stage of external review, sections of thesupporting documentation pertaining to human health wereconsidered by the following individuals, primarily to addressadequacy of coverage: J. Aquavella, Monsanto Company; M.Bird, Exxon Biomedical Sciences, Inc.; J.A. Bond, ChemicalIndustry Institute of Toxicology; I. Brooke, United KingdomHealth and Safety Executive; G. Granville, Shell Canada Ltd.;R. Keefe, Imperial Oil Ltd.; A. Koppikar, US EnvironmentalProtection Agency; R.J. Lewis, Exxon Biomedical Sciences, Inc.;K. Peltonen, Finnish Institute of Occupational Health; and F.Ratpan, Nova Chemicals

In the second stage of external review, accuracy ofreporting, adequacy of coverage, and defensibility ofconclusions with respect to hazard characterization andexposure–response analyses were considered in written review byBIBRA International and the following individuals: R.J. Albertini,University of Vermont; J.A. Bond, Chemical Industry Institute ofToxicology; I. Brooke, United Kingdom Health and SafetyExecutive; J. Bucher, US National Toxicology Program; B.Davis, US National Toxicology Program; E. Delzell, University ofAlabama at Birmingham; B.J. Divine, Texaco; A.A. Elfarra,University of Wisconsin-Madison; E. Frome, Oak Ridge NationalLaboratory; B.D. Goldstein, Environmental and OccupationalHealth Sciences Institute; R.F. Henderson, Lovelace RespiratoryResearch Institute; R.D. Irons, University of Colorado HealthSciences Center; A. Koppikar, US Environmental ProtectionAgency; J. Lubin, US National Cancer Institute; J. Lynch, ExxonBiomedical Sciences, Inc. (retired); R.L. Melnick, US NationalToxicology Program; K. Peltonen, Finnish Institute ofOccupational Health; A.G. Renwick, University of Southampton;


50

J. Siemiatycki, Institut Armand-Frappier; L.T. Stayner, USNational Institute for Occupational Safety and Health; J.A.Swenberg, University of North Carolina; R. Tice, IntegratedLaboratory Systems, Inc.; and J.B. Ward, Jr., University of TexasMedical Branch.

In the third and final stage of external expert review,adequacy of incorporation of the comments received during thesecond stage was considered at a final meeting of a panel of thefollowing members convened by Toxicology Excellence in RiskAssessment (TERA) in November 1998: H. Clewell, K.S. CrumpDivision of ICF Kaiser; M.L. Dourson, TERA; and L. Erdreich,Bailey Research Associates, Inc.

The health-related sections of the Assessment Reportwere reviewed and approved by the Health Protection BranchRisk Management meeting. The entire Assessment Report wasreviewed and approved by the Environment Canada/HealthCanada CEPA Management Committee.

Concurrent with review of the draft CICAD, there was alsoa public comment period for the source national assessment, inwhich the Priority Substances List Assessment Report was madeavailable for 60 days (2 October to 1 December 1999). Asummary of the comments and responses is available on theInternet at www.ec.gc.cceb1/eng/public/index_e.html.

APPENDIX 2 — CICAD PEER REVIEW

The draft CICAD on 1,3-butadiene was sent for review toinstitutions and organizations identified by IPCS after contactwith IPCS national Contact Points and Participating Institutions,as well as to identified experts. Comments were received from:

M. Baril, International Programme on Chemical Safety/Institut de Recherche en Santé et en Sécurité du Travaildu Québec, Canada

R. Benson, Drinking Water Program, US EnvironmentalProtection Agency, USA

T. Berzins, National Chemicals Inspectorate (KEMI), Sweden

R. Cary, Health and Safety Executive, United Kingdom

R. Chhabra, National Institute of Environmental HealthSciences, National Institutes of Health, USA

P. Edwards, Department of Health, United Kingdom

H. Gibb, National Center for Environmental Assessment,US Environmental Protection Agency, USA

R. Hertel, Federal Institute for Health Protection ofConsumers and Veterinary Medicine, Germany

J. Heuer, Federal Institute for Health Protection ofConsumers and Veterinary Medicine, Germany

J. Jinot, US Environmental Protection Agency, USA

C. Kimmel, US Environmental Protection Agency, USA

A.M. Koppikar, US Environmental Protection Agency,USA

S. Kristensen, National Industrial Chemicals Notificationand Assessment Scheme (NICNAS), Australia

N. Moore, BP Amoco Chemicals (commented throughDepartment of Health, United Kingdom)

H. Nagy, National Institute of Occupational Safety andHealth, USA

S. Tarkowski, Nofer Institute of Occupational Medicine,Poland

L. Vodickova, National Institute of Public Health, Centreof Industrial Hygiene and Occupational Diseases, CzechRepublic

P. Yao, Institute of Occupational Medicine, ChineseAcademy of Preventive Medicine, People’s Republic ofChina

K. Ziegler-Skylakakis, GSF-Forschungszentrum für Umveltund Gesundheit, Germany (transmitted comments fromBUA and industry representatives)


51

APPENDIX 3 — CICAD FINAL REVIEWBOARD

Helsinki, Finland, 26–29 June 2000

Members

Mr H. Ahlers, Education and Information Division, NationalInstitute for Occupational Safety and Health, Cincinnati, OH,USA

Dr T. Berzins, National Chemicals Inspectorate (KEMI), Solna,Sweden

Dr R.M. Bruce, Office of Research and Development, NationalCenter for Environmental Assessment, US EnvironmentalProtection Agency, Cincinnati, OH, USA

Mr R. Cary, Health and Safety Executive, Liverpool, UnitedKingdom (Rapporteur)

Dr R.S. Chhabra, General Toxicology Group, National Instituteof Environmental Health Sciences, Research Triangle Park, NC,USA

Dr H. Choudhury, National Center for Environmental Assessment,US Environmental Protection Agency, Cincinnati, OH, USA

Dr S. Dobson, Centre for Ecology and Hydrology, Monks Wood,Abbots Ripton, United Kingdom (Chairman)

Dr H. Gibb, National Center for Environmental Assessment, USEnvironmental Protection Agency, Washington, DC, USA

Dr R.F. Hertel, Federal Institute for Health Protection ofConsumers and Veterinary Medicine, Berlin, Germany

Ms K. Hughes, Priority Substances Section, EnvironmentalHealth Directorate, Health Canada, Ottawa, Ontario, Canada

Dr G. Koennecker, Chemical Risk Assessment, FraunhoferInstitute for Toxicology and Aerosol Research, Hanover,Germany

Ms M. Meek, Existing Substances Division, EnvironmentalHealth Directorate, Health Canada, Ottawa, Ontario, Canada

Dr A. Nishikawa, Division of Pathology, Biological SafetyResearch Centre, National Institute of Health Sciences, Tokyo,Japan

Dr V. Riihimäki, Finnish Institute of Occupational Health,Helsinki, Finland

Dr J. Risher, Agency for Toxic Substances and Disease Registry,Division of Toxicology, US Department of Health and HumanServices, Atlanta, GA, USA

Professor K. Savolainen, Finnish Institute of OccupationalHealth, Helsinki, Finland (Vice-Chairman)

Dr J. Sekizawa, Division of Chem-Bio Informatics, NationalInstitute of Health Sciences, Tokyo, Japan

Dr S. Soliman, Department of Pesticide Chemistry, Faculty ofAgriculture, Alexandria University, Alexandria, Egypt

Ms D. Willcocks, National Industrial Chemicals Notification andAssessment Scheme, Sydney, NSW, Australia

Observer

Dr R.J. Lewis (representative of European Centre forEcotoxicology and Toxicology of Chemicals), Epidemiology andHealth Surveillance, ExxonMobil Biomedical Sciences, Inc.,Annandale, NJ, USA

Secretariat

Dr A. Aitio, International Programme on Chemical Safety, WorldHealth Organization, Geneva, Switzerland (Secretary)

Dr P.G. Jenkins, International Programme on Chemical Safety,World Health Organization, Geneva, Switzerland

Dr M. Younes, International Programme on Chemical Safety,World Health Organization, Geneva, Switzerland


52

APPENDIX 4 — QUANTITATION OFEXPOSURE–RESPONSE FOR CRITICAL

EFFECTS ASSOCIATED WITH EXPOSURE TO1,3-BUTADIENE

Tumorigenic concentrations based onepidemiological data

Methods

The raw study data1 for the six plants investigated byDelzell et al. (1995) were used to calculate the potencyestimates. The data consisted of the cumulative occupationalexposures to butadiene and styrene at each year of eachsubject’s life (person-year), beginning with his entry into thecohort and terminating with death or other exit from the cohort.The data also contained information on race, age, calendaryear, and years since hire of each subject.

The response of interest was cases of death due to allforms of leukaemia, as information on the specific type ofleukaemia was insufficient; only cases in which leukaemia wasconsidered the underlying cause of death were considered inthese analyses. Exposure estimates were cumulativeoccupational exposures in ppm-years assumed to be incurred for8 h/day, 240 days/year over a 45-year working life.

The objective of this exposure–response analysis was tocompute the carcinogenic potency, expressed as the TC01, or theconcentration of butadiene associated with a 1% excessprobability of dying from leukaemia. This analysis involved twostages. First, the relationship between exposure and the deathrate due to leukaemia within the cohort was modelled. This wasaccomplished by collapsing (or stratifying) the data into discreteexposure categories and then modelling the mean exposure ineach category versus the death rates due to leukaemia. In thesecond stage of analysis, the TC01 was calculated based on thisexposure–response relationship and the background mortalityrates in the Canadian population.

Exposure–response modelling

In addition to stratifying by exposure, the data werestratified by race, age, calendar year, years since hire, andstyrene exposure in order to incorporate this information into theexposure–response relationship. Each of these variables wascollapsed into a small number of discrete categories in order toreduce the number of strata, thereby improving model stability.These variables and their categories are presented in Table A-1.Exposure, defined as the mean cumulative exposure per person-year, was calculated for person-years falling into each possiblecombination of the stratification variables.

The data were imported to Epicure (1993)2 for exposure–response modelling. All fitted models were of the form:

RR = O/E = g(D(t))

where RR is the rate ratio, O and E are the observed andexpected numbers of leukaemia deaths, D(t) is cumulativebutadiene exposure up to time t, and g is the exposure–responsemodel, which is constrained to pass through one at zeroexposure. Four different models, discussed in more detail below,were fitted to the data. At the model-fitting stage, the expectednumber of deaths is calculated on the basis of the non-exposedperson-years in the cohort, and not from population backgroundrates.

Lifetime probability of death due to leukaemia

Once the fitted exposure–response model was obtained,the lifetime probability of death due to leukaemia wascomputed using lifetable methods taking into account the deathrates in the Canadian population. The derivation of the formulaused for the lifetime probability of death due to leukaemiaproceeds as follows.

Let d(t) represent the exposure concentration of butadienein ppm at age t years, and let D(t) denote the cumulativeexposure in ppm-years, with:

D(t) = ∫t

dxxd0

)(

This formulation of cumulative exposure allows for the possibilityof non-constant exposure scenarios.

At a cumulative exposure of D(t) ppm-years, theprobability of dying from leukaemia by age t is given by:

(1)∫=t

R dxxSxxDhttDP0

)());(());((

where hR(D(t);t) is the mortality rate from leukaemia at age tgiven a cumulative exposure to butadiene of D(t), and S(t) is theprobability of survival up to age t. Equation (1) follows from theargument that the probability of death by age t is equal to theprobability of death at age t multiplied by the probability ofsurviving up until age t. In lifetable analysis, the mortality andsurvival rates are constant for each year, so the integral in (1)can be replaced by a summation over year.

Exposure to butadiene is assumed to augment the back-ground rate of leukaemia in a multiplicative fashion. In otherwords, the mortality rate, given exposure to butadiene, is equalto the background exposure rate multiplied by the excess riskdue to exposure to butadiene. This is known as the “proportionalhazard” model and is expressed as:

(2)[ ]))(()());(( tDgthttDH R ×=

where h(t) is the background mortality rate from leukaemia in thepopulation, calculated from Canadian age-specific death rates3

due to leukaemia, and g(D(t)) is the fitted exposure–responsemodel, or excess risk at age t.

The survival rate, S(t), appearing in equation (1) iscomputed from Canadian age-specific death rates due to allcauses, where the reported Canadian leukaemia mortality rate is

1 The cooperation of the sponsors and researchers for

the Delzell et al. (1995) study in the provision of thesedata is gratefully acknowledged.

2 Epicure is a collection of interactive programs used to

fit models to epidemiological data. The specific programused to model the data for this cohort of styrene-butadiene rubber workers is called AMFIT, which isspecially designed to model hazard functions forcensored cohort survival data. The strength of Epicurelies in its ability to easily allow the background rate todepend on user-specified strata, such as age, calendarperiod, and race.

3 Mortality data were provided to Health Canada by

Statistics Canada. The cooperation of the registrars ofvital statistics in the provinces and territories of Canadawho make mortality data available to Statistics Canadaunder federal–provincial agreements is gratefullyacknowledged.


53

Table A-1: Stratification variables for exposure–response modelling of epidemiological data from Delzell et al. (1995).

Variable Categories

Cumulative butadiene exposure (ppm-years) 0, >0–4, 5–9, 10–19, 20–29, 30–49, 50–99, 100–199, 200+

Cumulative styrene exposure (ppm-years) 0, >0–3, 4–6, 7–9, 10–19, 20–39, 40–59, 60–79, 80+

Race black, white, other

Age 40–44, 45–49, ..., 75–79, 80+

Calendar period 1940–44, 1945–49, ..., 1990–95

Years since hire 0–4, 5–9, …, 50–55

replaced by the modelled rate in order to incorporate exposureto butadiene. The formula describing the probability of survivalup to age i is given by:

(3)

+−−= ∑

=

i

jjjjji ghhhS

1

*exp

where h j* and h j are the Canadian mortality rates due to all

causes and due to leukaemia at age j, respectively, and g j =g(D(j)) is the excess risk at age j.

Substituting equation (2) into (1), the lifetime probabilityof death due to leukaemia is given by:

∑=

−=70

11)70);70((

iiii SghDP

where 1–70 years is the standard lifetime for a human.

Cancer potency (TC 01)

The TC01 is computed by determining the exposure D(t) atwhich the excess risk is equal to 0.01. That is,

01.0);0(1

);0());((=

−−

tPtPttDP

If a constant exposure d is assumed for an individual frombirth to age 70 years, then d(t) = d ppm and the cumulativeexposure D(t) = d × t ppm-years. The TC01 is then the ambientexposure level d (in ppm) at which the excess risk equals 0.01 att = 70 years.

Lagged exposure analysis

In separate analyses, exposures were lagged by n = 2, 5,10, 15, 20, and 25 years to determine if the models wouldprovide better fits if the most recent n years of exposure wereignored. An n-year lag was achieved by resetting an individual’scumulative exposure at each year to be equal to the exposurehe had accumulated n years prior. In so doing, the last n years ofexposure do not affect the probability of developing leukaemia.The data were first stratified on unlagged cumulative exposure,and then the individual exposures were lagged. Thus, thenumber of strata remains constant when using different lagperiods, and models with different lags may be directlycompared (Preston et al., 1987).

Validation study

To assess the predictive power of the exposure–responsemodels, a validation study was performed in which individuals inthe cohort were divided randomly into two groups. The modelswere fit separately to both groups, and then a likelihood ratiotest was performed to determine if the estimated parameterswere equal. The process of dividing and fitting was repeated

1000 times to characterize the variability due to the randomsplitting process. If the models provided consistent fits, then thelikelihood ratio test would be expected to reject at a rate equalto the desired significance level of the test (i.e., at a significancelevel of 0.05, the fitted parameters should be significantlydifferent 1 in 20 times). If the tests are significant more oftenthan this, the confidence in the predictive power of the models isreduced.

Results

Exposure–response modelling

Four different exposure–response models were examinedand are presented in Table A-2. These models are identical tothose fitted in the Delzell et al. (1995) report except that model2 is more general and flexible than the square root model usedby those authors. Preliminary analysis indicated that allstratification variables except race significantly affected themodel fit. Since race was only marginally insignificant, allvariables were used to stratify the data prior to model fitting.

The four models were fitted while stratifying on race, age,calendar year, years since hire, and styrene exposure. Theresults of the model fitting are displayed in Table A-2. (Note: Asmaller deviance roughly indicates a better fit.) A graphicrepresentation of the data and the fitted models is shown inFigure A-1. Judging from the model deviances and the shape ofthe curves relative to the data, especially in the low-dose region,model 1 provides the best fit to the data.

For purposes of comparison, the same models were fittedusing the median exposure as per the Delzell et al. (1995)report. These analyses indicated that there is little differencebetween using mean or median exposures. Models includingage as a multiplying factor of e(@age instead of as a stratificationvariable were also fitted, but these models did not fit as well.Since cumulative exposure and years since hire may beconfounded, their interaction was examined. The interactionwas not significant for any of the models. The same models wererefitted excluding the largest exposure group (200+ ppm-years),but this did not significantly affect any of the parameterestimates. The four models were also refitted allowing fordifferent background rates for control and exposed populations.Different background rates might be necessary in occupationalstudies where lifetime non-exposed workers may differfundamentally from exposed workers as a result of differences injobs and work areas. Results of this analysis indicated thatdifferent background rates are not necessary for these data.

The parameter estimates obtained in the present analysisare also not significantly different from those presented in theDelzell et al. (1995) report. The differences in parameterestimates are likely due to the different levels used in thestratification variables. Table A-2 compares the parameterestimates obtained in this analysis with those of the Delzell et al.(1995) report.


54

Table A-2: Parameter estimates and model deviances for each of four models fit to mean cumulative exposure per person-year forDelzell et al. (1995) study and comparison with parameter estimates from Delzell et al. (1995) analyses.

Model Parameterestimates Standard error Deviance

Parameter estimates fromDelzell et al. (1995) study p-value a

1) RR = (1 + dose)" " = 0.2850 SE(") = 0.0976 171.5 " = 0.2028 0.39

2) RR = 1 + $@dose" " = 0.3999$ = 0.4558

SE(") = 0.2733SE($) = 0.8222

172.0 " = 0.5000b

$ = 0.12930.62

3) RR = e$@dose $ = 0.0029 SE($) = 0.0014 176.7 $ = 0.0041 0.38

4) RR = 1 + $@dose $ = 0.0099 SE($) = 0.0065 174.7 $ = 0.0068 0.63a p-value of likelihood ratio test of equality of parameters.b For the Delzell et al. (1995) analysis, " was fixed at 0.5, and only $ was estimated.

Figure A-1. Observed rate ratios and fitted curves for leukaemia in Delzell et al. (1995) study.

Cancer potency (TC01)

The TC01s were calculated for each model using the lifetablemethods described above, and the resulting ambient occupationalexposures per person-year were converted to environmental exposuresby assuming that the exposures occurred for 8 h/day,240 days/year. This amounts to multiplying the TC01 by:

daysdays

hh

365

240

24

8×

To convert the ambient exposures from ppm to mg/m3, theTC01s are further multiplied by 2.21, the conversion factor forbutadiene. The occupational and equivalent environmental TC01s arepresented in Table A-3. Environmental TC01s for each of the fourmodels ranged from 1.4 to 4.3 mg/m3. TC01s calculated excluding thelargest exposure group were slightly smaller, ranging from 0.6 to 1.6mg/m3, while those calculated on the basis of median exposures weresimilar, ranging from 0.4 to 5.0 mg/m3.

TC01s were also calculated using the parameter estimates fromthe Delzell et al. (1995) report and are compared with the TC01sdeveloped here in Table A-3. They ranged from 3.1 to 14.3 mg/m3.

Lagged exposure analysis

The same four models were fitted when exposures were laggedby 2, 5, 10, 15, 20, and 25 years. The resulting model fits are displayedin Table A-4. Since the deviances are similar for each lag period, thisanalysis indicates that lagging exposures does not dramatically improvethe fit of any of the four models. In fact, TC01s for all four models andall lag periods ranged from 0.8 to 4.3 mg/m3.

Validation study

With respect to model validation, the p-values for the testsof equality of the parameters are displayed in Table A-5. If themodels were providing consistent fits between the two halves,the proportion of p-values less than the significance level of "would be ". The results of the simulation study indicate that thetest is rejecting more often than would be expected if the

Mean cumulative butadiene exposure per person year (ppm-years), adjusted for age, calendarperiod, race, years since hire and styrene exposure


55

Table A-3: Carcinogenic potency estimates (TC01s) for models fit to mean cumulative exposure per person-yearbased on Delzell et al. (1995) study and comparison with estimates from Delzell et al. (1995) analyses.

Current analysis Delzell et al. (1995) analysis

Model Occupational TC01

(mg/m3)Environmental TC 01

(mg/m3)Environmental TC 01

(mg/m3)

1) RR = (1 + dose)" 7.8 1.7 14.3

2) RR = 1 + $@dose" 6.5 1.4 6.4

3) RR = e$@dose 19.8 4.3 3.1

4) RR = 1 + $@dose 13.8 3.0 4.5

models were providing the same fits to both halves of the data.For model 1, the test was rejected at a significance level of 1%in 7.4% of the runs, whereas a rejection rate of 1% of the runswould be expected if the model was fitting consistently. Theresults of this analysis reduce the confidence in the power of themodels to predict leukaemia mortality rates.

Summary

It is noteworthy that the choice of the exposure–response modeldoes not have a large impact on the resulting TC01; as indicated in TableA-3, the values are similar, ranging from 1.4 to 4.3 mg/m3. However, ifa best model must be chosen, it would be model 1, owing to the smallerdeviance (Table A-2), the shape of the curve relative to the data in thelow-dose region (Figure A-1), and the fact that it has one fewerparameter than model 2, which provides a similar fit. The TC01 for model1 is 1.7 mg/m3.

It is difficult, though, to assess how well any of these modelstruly describes the data. It is noted that the plot in Figure A-1 providesonly a rough indication of the shape of the data, since each point on theplot is an average of data in many strata. The results of the validationstudy reduce confidence in the ability of the models to predictleukaemia mortality.

The choice of exposure lag does not greatly improve the fit ofany of the four models, and it does not affect the resulting TC01.Including all lagged models, the range of TC01s is still from 0.8 to 4.3mg/m3.

For comparison with these values, potency estimates were alsocalculated on the basis of the recent case–control study of styrene-butadiene rubber workers by Matanoski et al. (1997). Although workerswere from plants subsumed in the Delzell et al. (1995) study, exposurewas independently characterized. Treating the odds ratio presented bythese authors as a rate ratio (since leukaemia is a rare disease) andusing their model and parameter estimates as well as the same lifetablemethods described above, the TC01 for environmental exposure wascalculated to be 0.4 mg/m3. It is reassuring, therefore, that this value isonly slightly lower than the estimates derived on the basis of the Delzellet al. (1995) cohort study data.

Tumorigenic concentrations based on data fromstudies in experimental animals

Estimates of carcinogenic potency were calculated on the basisof the incidences of malignant lymphomas, histiocytic sarcomas,cardiac haemangiosarcomas, alveolar/bronchiolar adenomas orcarcinomas, hepatocellular adenomas or carcinomas, squamous cellpapillomas or carcinomas of the forestomach, adenomas or carcinomasof the Harderian gland, granulosa cell tumours of the ovaries, andadenoacanthomas, carcinomas, or malignant mixed tumours of themammary gland observed in B6C3F1 mice in the chronic bioassayconducted by the NTP (1993) and the mammary gland tumours,pancreatic exocrine adenomas, Leydig cell tumours, Zymbal glandcarcinomas, thyroid follicular cell adenomas or carcinomas, and uterinesarcomas in Sprague-Dawley rats reported by Hazleton LaboratoriesEurope Ltd. (1981a). (The tumour incidence data for each of the sitesconsidered are presented in Tables 2 and 3.)

In the NTP study, mice were exposed to 0, 6.25, 20, 62.5, 200,or 625 ppm (0, 13.8, 44.2, 138, 442, or 1383 mg/m3) butadiene for 6h/day, 5 days/week, for 103 weeks. Survival of mice decreased withincreasing exposure concentration; therefore, to minimize the effect ofthe high mortality rate, the poly-3 adjusted data (Bailer & Portier, 1988;Portier & Bailer, 1989) presented in the NTP (1993) report were used inthese calculations. For some tumour types, the adjusted data stilldemonstrated downward curvature at the highest concentration. In thesecases, the high-exposure group was excluded in the determination ofthe TC05. The TC05s were calculated for these end-points by first fittinga multistage model to the data. The multistage model is given by:

kk dqdqqedP −−−−−= ...101)(

where d is dose, k is the number of dose groups in the study minusone, P(d) is the probability of the animal developing a tumour at dose d,and qi > 0, i = 1,..., k are parameters to be estimated.

The models were fitted using GLOBAL82 (Howe & Crump, 1982),and a chi-square lack of fit test was performed for each model fit. Thedegrees of freedom for this test are equal to k minus the number of qi’swhose estimates are non-zero. A p-value less than 0.05 indicates asignificant lack of fit. The lower confidence limits presented areapproximate, based on output from GLOBAL82. Results from the modelfitting are displayed in Table A-6. Plots of the data and the fittedmodels are shown in Figure A-2.

TC05s were determined as the doses D (in mg/m3) that satisfy

05.0)0(1

)0()(=

−−

PPDP

and then adjusted by multiplying by:

2

104104/7/5

/24/6

weeksweeksw

weeksweeksw

weekdaysweekdays

dayhdayh

×××

where, in the first term, which amortizes the dose to be constant overthe lifetime of a mouse, w is the duration of the experiment (103weeks). The second factor was suggested by Peto et al. (1984) andcorrects for an experiment length that is unequal to the standardlifetime. Since tumours develop much more rapidly later in life, agreater than linear increase in the tumour rate is expected when animalsare observed for tumours longer than their standard lifetime (or thereverse, when animals are observed for a period shorter than theirstandard lifetime). (Application of this factor does not impact greatly onthe final values, since it is very close to one.) The selected TC05 valuesfor this study and their 95% lower confidence limits (LCLs) arepresented in Table A-6 and range from 2.3 mg/m3 (95% LCL = 1.7mg/m3) or 1.1 ppm (95% LCL = 0.79 ppm) for Harderian gland tumoursin males to 99 mg/m3 (95% LCL = 23 mg/m3) or 45 ppm (95% LCL = 10ppm) for malignant lymphomas in males.


56

Table A-4: Parameter estimates and model deviances for each of four lagged-exposure models fitted to median cumulative exposure per person-year.

Model Lag Parameter estimates Standard error Deviance

1) RR = (1 + dose)" None " = 0.2850 SE(") = 0.0976 171.5

2 years " = 0.2852 SE(") = 0.0982 171.6

5 years " = 0.2883 SE(") = 0.0995 171.6

10 years " = 0.3064 SE(") = 0.1034 171.1

15 years " = 0.2955 SE(") = 0.1079 172.4

20 years " = 0.2891 SE(") = 0.1141 173.6

25 years " = 0.2898 SE(") = 0.1334 175.4

2) RR = 1 + $@dose" None " = 0.3999 SE(") = 0.2733 172.0

$ = 0.4557 SE($) = 0.8219

2 years " = 0.3992 SE(") = 0.2739 172.0

$ = 0.4602 SE($) = 0.8279

5 years " = 0.4024 SE(") = 0.2737 172.0

$ = 0.4647 SE($) = 0.8288

10 years " = 0.4245 SE(") = 0.2755 171.4

$ = 0.4693 SE($) = 0.8345

15 years " = 0.4835 SE(") = 0.3397 172.6

$ = 0.2878 SE($) = 0.5846

20 years " = 0.4720 SE(") = 0.3558 173.9

$ = 0.3243 SE($) = 0.6572

25 years " = 0.2960 SE(") = 0.2833 175.3

$ = 0.9293 SE($) = 1.5710

3) RR = e$@dose None $ = 0.0029 SE($) = 0.0014 176.7

2 years $ = 0.0029 SE($) = 0.0015 176.8

5 years $ = 0.0031 SE($) = 0.0015 176.7

10 years $ = 0.0034 SE($) = 0.0016 176.4

15 years $ = 0.0035 SE($) = 0.0018 177.0

20 years $ = 0.0033 SE($) = 0.0022 178.2

25 years $ = 0.0033 SE($) = 0.0022 178.2

4) RR = 1 + $@dose None $ = 0.0099 SE($) = 0.0065 174.7

2 years $ = 0.0102 SE($) = 0.0067 174.7

5 years $ = 0.0109 SE($) = 0.0072 174.6

10 years $ = 0.0137 SE($) = 0.0089 173.8

15 years $ = 0.0158 SE($) = 0.0106 174.1

20 years $ = 0.0179 SE($) = 0.0129 175.7

25 years $ = 0.0179 SE($) = 0.0129 175.7

Estimates of carcinogenic potency were also calculatedbased on the results of the bioassay in Sprague-Dawley rats(Hazleton Laboratories Europe Ltd., 1981a). In this study, ratswere exposed to 0, 1000, or 8000 ppm (0, 2212, or 17 696mg/m3) for 6 h/day, 5 days/week, for 105 (males) or 111 (females)weeks. A high mortality rate was observed at the higherconcentration; therefore, this exposure group was excluded fromthe analysis, except for the potency estimates for pancreaticexocrine adenomas in males (for this end-point, exclusion of thehigh-exposure group would have resulted in theexposure–response relationship curving downwards). As for mice,

a multistage model was fit to the data for rats using GLOBAL82and adjusted to account for study duration (w) by multiplying by:

2

104104/7/5

/24/6

weeksweeksw

weeksweeksw

weekdaysweekdays

dayhdayh

×××

where the duration of the experiment was 105 weeks for malesand 111 weeks for females. The exposure–response curves andestimated adjusted TC05 values based on this study in rats arepresented in Figure A-3 and Table A-6, respectively. The


57

Table A-5: Model validation p-values for Delzell et al. (1995) study.

Proportion of p-valuesa that are

Model less than 0.01 less than 0.05 less than 0.1

1) RR = (1 + dose)" 0.074 0.167 0.252

2) RR = 1 + $@dose" 0.084 0.19 0.286

3) RR = e$@dose 0.08 0.188 0.264

4) RR = 1 + $@dose 0.103 0.214 0.303

a p-value of likelihood ratio test of equality of parameters fitted to each half of the data.

concentrations of butadiene estimated to be associated with a5% increased incidence of tumours ranged from 6.7 mg/m3 (95%LCL = 4.7 mg/m3) or 3.0 ppm (95% LCL = 2.1 ppm) to 4872mg/m3 (95% LCL = 766 mg/m3) or 2203 ppm (95% LCL = 346ppm) for tumours of the mammary gland and Zymbal gland infemale rats, respectively. Although the available data foranalysis of exposure–response were more limited for rats than formice, it is interesting to note the similarity in estimates ofpotency for mammary gland tumours (i.e., 6.7 mg/m3 in bothspecies).

Based on modelling (using THC; Howe, 1995a) of theincidence of micronucleated polychromatic erythrocytes inB6C3F1 mice exposed to butadiene for up to 15 months in theNTP bioassay, BMC05s for somatic cell mutations were verysimilar to the lower end of the range of the TC05s for tumourinduction.

Benchmark concentrations for ovarian atrophyin mice

Benchmark concentrations (BMC05s) for ovarian atrophywere derived on the basis of the chronic bioassay conducted bythe NTP (1993) in which B6C3F1 mice were exposed toconcentrations of 0, 6.25, 20, 62.5, 200, or 625 ppm (0, 13.8,44.2, 138, 442, and 1383 mg/m3) butadiene for up to 2 years.There was a concentration-related increase in the incidence aswell as the severity of ovarian atrophy, as summarized in TableA-7.

The exposure–response relationship for ovarian atrophyfrom this study was quantified by fitting the following model tothe dose–response data (Howe, 1995b):

[ ]

≤−⋅−+≤=

−−−−−0

)(...)(00

00

if1)1(

if)(001 ddeqq

ddqdP kk ddqddq

where d is dose, k is the number of dose groups in the studyminus one, P(d) is the probability of the animal developing theeffect at dose d, and q i > 0, i =1,..., k and d0 are parameters to beestimated. The models were fit using THRESH (Howe, 1995b),and the BMC05s were calculated as the dose D that satisfies:

05.0)0(1

)0()(=

−−

PPDP

A chi-square lack of fit test was performed for each of themodel fits. The degrees of freedom for this test are equal to kminus the number of q i’s whose estimates are non-zero. A p-value less than 0.05 indicates a significant lack of fit.

The BMC05 was then amortized to be constant over thestandard life of a mouse by multiplying by:

weekdaysweekdays

dayhdayh

/7

/5

/24

/6×

Resulting BMC05s and lack of fit information for all models fit aredisplayed in Table A-8.

The model fitted to all six exposure groups exhibited asignificant lack of fit, likely due to the fact that the curve risessharply and then plateaus at the three highest exposure groups.Plots of the data and fitted model are displayed in Figure A-4.Since a good fit in the range of the BMC05 (in the vicinity of 6.25ppm [13.8 mg/m3]) is desired, the model was refitted omittingthe two highest exposure groups. This model again indicates amarginal lack of fit. The graph of this model (Figure A-5)indicates that this model provides a reasonable visual fit to thedata, but the resulting BMC05 is uncertain due to lack of fit of themodel.

The BMC05 for the model excluding the two highest dosegroups was calculated to be 0.57 mg/m3, with a 95% LCL of0.44 mg/m3.

If only those animals that had moderate or markedovarian atrophy from all exposure groups were included, theresulting BMC05 would be 9.6 mg/m3 (95% LCL = 7.6 mg/m3),although there is again a significant lack of fit (Figure A-6). If thehighest exposure group is excluded, the BMC05 for moderate ormarked ovarian atrophy becomes 3.1 mg/m3, with a 95% LCL of2.5 mg/m3 (Figure A-7).

Table A-6: Carcinogenic potency estimates (TC05s)a of butadiene based on results of bioassays in experimental animals.

Males Females

Tumour typeTC05

(mg/m3)95%LCL

(mg/m3)PP2 b dfc pd

TC05

(mg/m3)95% LCL(mg/m3) PP2 b dfc pd

Mice (from NTP, 1993)

Alveolar/bronchiolar adenomas or carcinomas 2.4 1.4 1.0 3 0.79 5.2 3.2 9.1 4 0.06

Histiocytic sarcomas 12 8.4 7.6 5 0.18 21 12 5.4 4 0.25

Cardiac haemangiosarcomas 14 6.4 0.34 3 0.95 7.6 5.2 18 4 0.00

Forestomach squamous cell papillomas or carcinomas 29 13 6.1 3 0.11 14 8.1 4.3 4 0.36

Ovarian granulosa cell tumours – – – – – 6.7 4.4 5.0 3 0.17

Mammary gland adenoacanthomas, carcinomas, or malignant mixedtumours

– – – – – 6.7 4.9 13 4 0.01

Hepatocellular adenomas or carcinomas 3.2 1.9 2.8 2 0.24 5.4 3.2 0.8 3 0.85

Harderian gland adenomas or carcinomas 2.3 1.7 0.5 2 0.77 4.7 2.7 1.5 2 0.47

Malignant lymphomase 99 23 3.3 3 0.35 23 6.9 3.9 3 0.27

Rats (from Hazleton Laboratories Europe Ltd., 1981a)

Mammary gland adenomas or carcinomas – – – – – 6.7 4.7 0 0 –

Pancreatic exocrine adenomas 597 316 1.1 1 0.29 – – – – –

Leydig cell tumours 161 96 0 1 – – – – – –

Thyroid follicular cell adenomas or carcinomas – – – – – 142 113 0 1 –

Uterine sarcomas – – – – – 189 113 0 0 –

Zymbal gland carcinomas 1023 905 1 1 0.32 4872 766 0.06 2 0.97

a Values have been adjusted for lifetime exposure.b Chi-squared goodness of fit statistic.c Degrees of freedom.d p-value of goodness of fit test (p-value < 0.05 indicates significant lack of fit).e Values for malignant lymphomas presented here only for comparison; potency estimates for these tumours not considered relevant to humans due to the greater sensitivity of these mice to induction of this

effect associated with the presence of an endogenous retrovirus.

59

Figure A-2: Exposure-response analysis for butadiene-induced tumours in mice

60

Figure A-2 (continued).

61

Fiure A2 Continued).

62

Fig A-2 (continued).

63

Figure A-3: Exposure-response analysis for butadiene-induced tumours in rats

64

Figure A-3 (continued).


65

Table A-7: Incidence and severity of ovarian atrophy observed in 2-year bioassay in mice (NTP, 1993).

Exposure level(ppm)

Number of animalsexamined

All severities

Minimalseverity

Mildseverity

Moderateseverity

Markedseverity

0 49 4 1 2 1 0

6.25 49 19 0 15 4 0

20 48 32 1 23 8 0

62.5 50 42 3 18 21 0

200 50 43 0 9 34 0

625 79 69 0 19 47 3

Table A-8: Benchmark concentrations for ovarian atrophy.

Ovarian atrophy BMC05

(ppm)

95% LCL onBMC05

(ppm)BMC05

(mg/m3)

95% LCL onBMC05

(mg/m3)Chi-

square df p-value

All severities 2.5 1.9 5.6 4.1 61 4 0.00

All severities, excludingtop two dose groups

0.25 0.20 0.57 0.44 7.0 2 0.03

Moderate/markedseverity

4.3 3.4 9.6 7.6 37.1 4 0.00

Moderate/markedseverity, excluding topdose group

1.4 1.1 3.1 2.5 2.2 3 0.55


66

Figure A-4: Exposure–response analysis for ovarian atrophy in mice(* BMC05 and BMCL05 unadjusted for lifetime dosing).

Figure A-5: Exposure–response analysis for ovarian atrophy in mice excluding two highest dose groups(* BMC05 and BMCL05 unadjusted for lifetime dosing).


67

Figure A-6: Exposure–response analysis for moderate/marked ovarian atrophy(* BMC05 and BMCL05 unadjusted for lifetime dosing).

Figure A-7: Exposure–response analysis for moderate/marked ovarian atrophy, excluding high-dose group(* BMC05 and BMCL05 unadjusted for lifetime dosing).

Prepared in the context of cooperation between the InternationalProgramme on Chemical Safety and the European Commission

© IPCS 2000

SEE IMPORTANT INFORMATION ON THE BACK.

IPCSInternationalProgramme onChemical Safety

1,3-BUTADIENE 0017April 2000

CAS No: 106-99-0RTECS No: EI9275000UN No: 1010 (stabilized)EC No: 601-013-00-X

DivinylVinylethyleneC4H6 / CH2=(CH)2=CH2

Molecular mass: 54.1

TYPES OFHAZARD/EXPOSURE

ACUTE HAZARDS/SYMPTOMS PREVENTION FIRST AID/FIRE FIGHTING

FIRE Extremely flammable. NO open flames, NO sparks, andNO smoking.

Shut off supply; if not possible andno risk to surroundings, let the fireburn itself out; in other casesextinguish with water spray, powder,carbon dioxide.

EXPLOSION Gas/air mixtures are explosive. Closed system, ventilation,explosion-proof electrical equipmentand lighting. Prevent build-up ofelectrostatic charges (e.g., bygrounding) if in liquid state.

In case of fire: keep cylinder cool byspraying with water.

EXPOSURE AVOID ALL CONTACT! AVOIDEXPOSURE OF (PREGNANT)WOMEN!

Inhalation Cough. Sore throat. Dizziness.Headache. Drowsiness. Blurredvision. Nausea. Unconsciousness.

Ventilation, local exhaust, orbreathing protection.

Fresh air, rest. Refer for medicalattention.

Skin ON CONTACT WITH LIQUID:FROSTBITE.

Cold-insulating gloves. ON FROSTBITE: rinse with plentyof water, do NOT remove clothes.Refer for medical attention.

Eyes Redness. Pain. Blurred vision. SeeSkin.

Safety goggles. First rinse with plenty of water forseveral minutes (remove contactlenses if easily possible), then taketo a doctor.

Ingestion Do not eat, drink, or smoke duringwork.

SPILLAGE DISPOSAL PACKAGING & LABELLING

Evacuate danger area! Consult an expert!Ventilation. NEVER direct water jet on liquid.Remove all ignition sources. Chemical protectionsuit including self-contained breathing apparatus.

F+ SymbolT SymbolR: 45-12S: 53-45Note: DUN Hazard Class: 2.1

Do not transport with food andfeedstuffs.

EMERGENCY RESPONSE STORAGE

Transport Emergency Card: TEC (R)-20G41NFPA Code: H2; F4; R2

Fireproof. Cool. Separated from food and feedstuffs.

Boiling point: -4°CMelting point: -109°CRelative density (water = 1): 0.6Solubility in water: none (0.1 g/100 ml)Vapour pressure, kPa at 20°C: 245

Relative vapour density (air = 1): 1.9 3,47 (1.87)Flash point: -76°CAuto-ignition temperature: 414°CExplosive limits, vol% in air: 1.1-16.3Octanol/water partition coefficient as log Pow: 1.99

LEGAL NOTICE Neither the EC nor the IPCS nor any person acting on behalf of the EC or the IPCS is responsible for the use which might be made of this information

©IPCS 2000

0017 1,3-BUTADIENE

IMPORTANT DATA

Physical State; AppearanceCOLOURLESS COMPRESSED LIQUEFIED GAS, WITHCHARACTERISTIC ODOUR.

Physical dangersThe gas is heavier than air and may travel along the ground;distant ignition possible. As a result of flow, agitation, etc.,electrostatic charges can be generated. The vapours areuninhibited and may form polymers in vents or flame arrestersof storage tanks, resulting in blockage of vents.

Chemical dangersThe substance can under specific circumstances (exposure toair) form peroxides, initiating explosive polymerization. Thesubstance may polymerize due to warming with fire or explosionhazard. Shock-sensitive compounds are formed with copperand its alloys (see Notes). The substance decomposesexplosively on rapid heating under pressure. Reacts vigorouslywith oxidants and many other substances, causing fire andexplosion hazard.

Occupational exposure limitsTLV: (as TWA) 2 ppm; A2 (ACGIH 1999).MAK: class 2 (1999)

Routes of exposureThe substance can be absorbed into the body by inhalation.

Inhalation riskA harmful concentration of this gas in the air will be reachedvery quickly on loss of containment.

Effects of short-term exposureThe substance irritates the eyes and the respiratory tract. Rapidevaporation of the liquid may cause frostbite. The substancemay cause effects on the central nervous system, resulting inlowering of consciousness.

Effects of long-term or repeated exposureThe substance may have effects on the bone marrow, resultingin leukemia. This substance is probably carcinogenic tohumans. May cause heritable genetic damage in humans.Animal tests show that this substance possibly causes toxiceffects upon human reproduction.

PHYSICAL PROPERTIES

ENVIRONMENTAL DATA

NOTES

Piping material for this gas must not contain over 63% of copper.Use of alcoholic beverages enhances the harmful effect.The odour warning when the exposure limit value is exceeded is insufficient.

ADDITIONAL INFORMATION


70

RÉSUMÉ D’ORIENTATION

Ce CICAD sur le 1,3-butadiène a été préparé par laDirection de l’Hygiène de l’Environnement de SantéCanada sur la base d’une documentation préparéesimultanément dans le cadre du Programme d’évaluationdes substances prioritaires, en application de la Loicanadienne sur la protection de l’environnement (LCPE).Les évaluations sanitaires des substances prioritaireseffectuées en application de cette loi portent sur leseffets potentiels que ces produits pourraient avoir sur lasanté humaine en cas d’exposition indirecte dansl’environnement général. Cette mise au point prend encompte des données allant jusqu’à fin avril 1998.L’appendice 1 donne des informations sur la nature del’examen par des pairs ainsi que sur les sourcesdocumentaires utilisées. Des renseignements surl’examen de ce CICAD par des pairs sont donnés àl’appendice 2. Ce CICAD a été approuvé en tantqu’évaluation internationale lors de la réunion du Comitéd’évaluation finale qui s’est tenue à Helsinki (Finlande)du 26 au 29 juin 2000. La liste des participants à cetteréunion figure à l’appendice 3. La fiche internationale surla sécurité chimique (ICSC 0017) du 1,3-butadiène, établiepar le Programme international sur la sécurité chimique(IPCS, 1993) est également reproduite dans le présentdocument.

Le 1,3-butadiène (No CAS 106-99-0) résulte de lacombustion incomplète de certains produits lors deprocessus naturels ou d’activités humaines. C’estégalement un produit chimique industriel principalementutilisé pour la préparation de polymères, notamment dupolybutadiène et des caoutchoucs et latex styrène-butadiène ou des élastomères acrylonitrile-butadiène. Le1,3-butadiène pénètre dans l’environnement avec les gazd’échappement des véhicules à essence ou à gazole, lorsde l’utilisation de combustibles à d’autres fins que letransport, de la combustion de la biomasse ou encorelors de son utilisation sur des sites industriels.

Le 1,3-butadiène ne persiste pas dans l’environne-ment mais il est néanmoins répandu dans tout le milieuurbain du fait de la présence généralisée des sources decombustion où il prend naissance. C’est dans les villeset à proximité immédiate des installations industriellesque sont mesurées les concentrations atmosphériquesles plus fortes.

L’exposition de la population générale au 1,3-butadiène est due principalement à sa présence dans l’airextérieur ou intérieur. Comparativement, la contributiondes autres véhicules, comme les aliments ou l’eau deboisson, reste négligeable. En revanche, celle de lafumée de tabac peut être importante.

Le métabolisme du 1,3-butadiène se révèle être demême nature d’une espèce à l’autre, mais il y a néan-

moins des différences quantitatives notamment en ce quiconcerne la proportion de métabolites présumés toxiquesqui se forment; ainsi chez la souris le métabolismeoxydatif en mono-, puis en diépoxyde est plus importantque chez le rat ou l’Homme. Par ailleurs, il peut égalementy avoir des variations interindividuelles dans lacapacicité de métabolisation chez l’Homme, variationsqui s’expliquent par le polymorphisme génétique desenzymes en cause.

Le 1,3-butadiène présente une faible toxicité aiguëchez les animaux de laboratoire. On a toutefois observéune atrophie ovarienne à toutes les concentrations chezdes souris exposées à ce composé pendant une périodeprolongée. D’autres effets sur les ovaires ont égalementété relevés lors d’études à court terme. Chez les sourismâles, on a constaté que le 1,3-butadiène provoquait uneatrophie testiculaire, mais à des concentrations supéri-eures à celles qui étaient toxiques pour les femelles. Lesdonnées limitées dont on dispose ne permettent pas deconclure que ce composé est tératogène pour la progéni-ture des animaux de laboratoire mâles ou femelles qui yont été exposés, ni qu’il est véritablement toxique pour lefoetus à des concentrations inférieures à celles qui sonttoxiques pour la mère.

Le 1,3-butadiène exerce également divers effets surle sang et la moelle osseuse de la souris; en ce qui con-cerne le rat, les données, bien que limitées, ne mettentpas en évidence d’effets de ce genre.

L’ensemble des études montre qu’une fois inhalé,le 1,3-butadiène se révèle fortement cancérogène chez lasouris, provoquant des tumeurs de localisationsmultiples à toutes concentrations. Il se révèle égalementcancérogène chez le rat à toutes les concentrations,selon la seule étude adéquate dont on dispose. Chezcette espèce, les seules concentrations étudiées étaientbeaucoup plus fortes que pour la souris, mais il apparaîtnéanmoins que le rat est l’espèce la moins sensible, sil’on prend comme élément de comparaison l’incidencedes tumeurs. Il est probable que la plus grandesensibilité observée chez les souris est liée auxdifférences interspécifiques mentionnées plus hautconcernant le métabolisme, et notamment le métabolismequi conduit à la formation d’époxydes actifs.

Le 1,3-butadiène a une action mutagène sur lescellules somatiques de la souris et du rat, avec uneactivité plus forte chez la souris. On a égalementconstaté d’autres lésions génétiques dans les cellulessomatiques de cette dernière espèce, à l’exclusion decelles du rat. Le composé s’est révélé systématiquementgénotoxique pour les cellules germinales de la souris,mais apparemment pas pour celles du rat, selon la seuleétude retrouvée.

Toutefois il n’y a apparemment pas de différenceinterspécifique pour ce qui est de la sensibilité aux effets


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génétiques dus aux époxydes résultant de la métabolisa-tion du 1,3-butadiène. Selon certaines données con-cernant l’exposition professionnelle mais qui restenttoutefois limitées, le 1,3-butadiène serait égalementgénotoxique pour l’Homme et provoquerait des lésionsmutagènes et clastogènes dans les cellules somatiques.

Il existe, entre l’exposition au 1,3-butadiène sur lelieu de travail et certaines leucémies, une corrélation quirépond à plusieurs des critères d’une relation de cause àeffet. L’étude la plus vaste et la plus exhaustiveeffectuée jusqu’ici, et qui a porté sur une cohorte detravailleurs employés dans de multiples usines, a mis enévidence un parallélisme entre une augmentationconstatée de la mortalité par leucémie et l’expositioncumulée estimative au 1,3-butadiène dans les industriesproduisant des élastomères styrène-butadiène. Lacorrection pour tenir compte de l’exposition au benzèneet au styrène n’a pas fait disparaître cette corrélation, quise manifestait d’ailleurs le plus fortement dans les sous-groupes potentiellement les plus exposés. On aégalement observé une corrélation entre l’exposition au1,3-butadiène et les leucémies lors d’une étude cas-témoins menée indépendamment de l’enquête précitéesur une population de travailleurs en grande partieidentique. Par contre, on n’a pas constatéd’augmentation de la mortalité par leucémie chez destravailleurs employés à la production de butadiènemonomère mais non exposés simultanément à certainesdes autres substances présentes dans l’industrie desélastomères styrène-butadiène, en dépit d’élémentsd’appréciation limités concernant la possibilité d’uneassociation avec la mortalité par lymphosarcomes ouréticulosarcomes dans certains sous-groupes.

Au vu des données épidémiologiques et toxicolo-giques, le 1,3-butadiène est cancérogène pour l’Hommeet pourrait également être génotoxique. Le pouvoircancérogène (concentration qui entraîne une augmen-tation de 1% de la mortalité par leucémie) a été évalué à1,7 mg/m3 en s’appuyant sur les résultats de l’enquêteépidémiologique la plus vaste et la mieux conduite surdes travailleurs exposés. Cette valeur correspond àl’extrémité inférieure de la série de concentrationstumorigènes déterminée par des études sur desrongeurs. Le 1,3-butadiène présente également unetoxicité génésique chez les animaux de laboratoire. Cepotentiel toxique peut s’évaluer par la concentration deréférence de 1,3-butadiène obtenue dans le cas des effetsovariens et qui est égale à 0,57 mg/m3.

Les effets sanitaires du 1,3-butadiène et le moded’action de ce composé ont été étudiés en détail, maisd’importants travaux de recherche continuent de lui êtreconsacrés afin de tenter de lever les incertitudes quisubsistent dans la base de données.

RESUMEN DE ORIENTACIÓN

Este CICAD sobre el 1,3-butadieno, preparado porla Dirección General de Higiene del Medio del Ministeriode Salud del Canadá, se basó en la documentaciónpreparada al mismo tiempo como parte del Programa deSustancias Prioritarias en el marco de la Ley Canadiensede Protección del Medio Ambiente (CEPA). El objetivode las evaluaciones sanitarias de las sustancias priori-tarias en el marco de la CEPA es valorar los efectospotenciales de la exposición indirecta en el medioambiente general para la salud humana. En este examense utilizaron los datos obtenidos hasta el final de abril de1998. La información relativa al carácter del examencolegiado y a la disponibilidad del documento original sepresenta en el apéndice 1. La información sobre elexamen colegiado de este CICAD figura en el apéndice 2.Este CICAD se aprobó como evaluación internacional enuna reunión de la Junta de Evaluación Final, celebradaen Helsinki (Finlandia) del 26 al 29 de junio de 2000. Lalista de participantes en esta reunión figura en elapéndice 3. La Ficha internacional de seguridad química(ICSC 0017) para el 1,3-butadieno, preparada por elPrograma Internacional de Seguridad de las SustanciasQuímicas (IPCS, 1993), también se reproduce en elpresente documento.

El 1,3-butadieno (CAS Nº 106-99-0) se produce poruna combustión incompleta en procesos naturales yactividades humanas. Es también un producto químicoindustrial que se utiliza fundamentalmente en laproducción de polímeros, en particular polibutadieno,cauchos y látex de estireno-butadieno y cauchos denitrilo-butadieno. El 1,3-butadieno se incorpora al medioambiente a partir de las emisiones de los gases de escapede los vehículos con motor de gasolina y diésel, de lacombustión de combustibles fósiles no utilizados en eltransporte, de la combustión de biomasa y de usosindustriales sobre el terreno.

El 1,3-butadieno no es persistente, pero sí ubicuoen el medio ambiente urbano, debido a la presenciageneralizada de sus fuentes de combustión. Lasconcentraciones atmosféricas más altas se han medidoen el aire de las ciudades y en las cercanías de fuentesindustriales.

La población general está expuesta al 1,3-buta-dieno fundamentalmente mediante el aire del ambiente yde los espacios cerrados. En comparación, otros medios,entre ellos los alimentos y el agua de bebida, contrib-uyen de manera insignificante a la exposición a él. Elhumo del tabaco puede producir cantidades significa-tivas de1,3-butadieno.

El metabolismo del 1,3-butadieno parece sercualitativamente semejante en todas las especies,aunque hay diferencias cuantitativas en los metabolitos


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supuestamente tóxicos que se forman; los ratonesparecen oxidar el 1,3-butadieno a monoepóxido ydespués a diepóxido en mayor medida que las ratas o laspersonas. Sin embargo, la capacidad metabólica para el1,3-butadieno puede variar de unas personas a otras, enfunción del polimorfismo genético de las enzimascorrespondientes.

El 1,3-butadieno tiene una toxicidad aguda baja enanimales experimentales. Sin embargo, la exposiciónprolongada de ratones a este producto se asoció con laaparición de atrofia ovárica con todas las concentra-ciones utilizadas en los ensayos. En estudios de expo-sición más breve se observaron también efectos en losovarios. En los ratones machos se detectó asimismoatrofia testicular a concentraciones superiores a lasasociadas con efectos en las hembras. Sobre la base delos limitados datos disponibles, no hay pruebas conclu-yentes de que el 1,3-butadieno sea teratogénico en losanimales experimentales tras la exposición materna opaterna o de que induzca una toxicidad fetal importante aconcentraciones inferiores a las que provocan toxicidadmaterna.

El 1,3-butadieno indujo también diversos efectosen la sangre y en la médula ósea de ratones; aunque losdatos son limitados, no se han observado efectos seme-jantes en las ratas.

El 1,3-butadieno inhalado es un potente carcinó-geno en ratones, induciendo tumores en lugaresmúltiples a todas las concentraciones utilizadas en todoslos estudios identificados. Fue también carcinogénico enratas a todos los niveles de exposición en el únicoestudio disponible al respecto; aunque en las ratas sólose ensayaron concentraciones mucho más altas que enlos ratones, las ratas parecen ser la especie menossensible, tomando como base la comparación de losdatos sobre la incidencia de tumores. La mayorsensibilidad de los ratones que las ratas a la inducciónde estos efectos por el 1,3-butadieno probablemente sedebe a diferencias entre las especies con respecto almetabolismo de formación de epóxidos activos.

El 1,3-butadieno es mutagénico en célulassomáticas tanto de ratones como de ratas, aunque lapotencia mutagénica fue mayor en los primeros. De lamisma manera, el 1,3-butadieno indujo otros dañosgenéticos en células somáticas de ratones, pero no enlas de ratas. Fue también sistemáticamente genotóxico encélulas germinales de ratones, pero no en la únicavaloración identificada en ratas. Sin embargo, no seobservaron diferencias aparentes en la sensibilidad delas especies a los efectos genéticos inducidos por losmetabolitos epóxidos del 1,3-butadieno. Hay tambiénpruebas limitadas de poblaciones expuestas en el lugarde trabajo de que el 1,3-butadieno es genotóxico para laspersonas, induciendo daños mutagénicos yclastogénicos en las células somáticas.

La asociación entre la exposición al 1,3-butadienoen el entorno de trabajo y la leucemia cumple varios delos criterios tradicionalmente establecidos para lacausalidad. En el estudio más amplio y completorealizado hasta el momento, utilizando una cohorte detrabajadores de diversas instalaciones, la mortalidad porleucemia aumentó con una exposición acumulativaestimada al 1,3-butadieno en la industria del caucho deestireno-butadieno; esta asociación se mantuvo tras elcontrol de la exposición al estireno y al benceno yalcanzó la intensidad máxima en los subgrupos con elmayor potencial de exposición. Análogamente, en unestudio de casos y testigos independiente basado prác-ticamente en la misma población de trabajadores seobservó una asociación entre la exposición al 1,3-buta-dieno y la leucemia. Sin embargo, no se produjo unaumento de la mortalidad a causa de la leucemia en lostrabajadores de la producción de monómeros de buta-dieno que no estaban simultáneamente expuestos aalgunas de las otras sustancias presentes en la industriadel caucho de estireno-butadieno, aunque en algunossubgrupos se obtuvieron pruebas limitadas deasociación con la mortalidad por linfosarcoma yreticulosarcoma.

Los datos epidemiológicos y toxicológicos dispon-ibles demuestran que el 1,3-butadieno es carcinogénico,y también puede ser genotóxico, para las personas. Secalculó una potencia carcinogénica (concentración queproduce un aumento del 1% de la mortalidad por leu-cemia) de 1,7 mg/m3, sobre la base de los resultados de lamayor investigación epidemiológica bien realizada entrabajadores expuestos. Este valor es semejante al nivelmás bajo de la gama de concentraciones tumorígenasdeterminadas a partir de los estudios realizados conroedores. El 1,3-butadieno indujo también toxicidadreproductiva en animales experimentales. Como medidade su potencia para inducir efectos reproductivos, seobtuvo una concentración de referencia de 0,57 mg/m3

para la toxicidad ovárica en los ratones.

Aunque se han investigado a fondo los efectos enla salud asociados con la exposición al 1,3-butadieno y elmecanismo de acción para la inducción de estos efectos,se siguen realizando numerosas investigaciones sobreesta sustancia, a fin de tratar de abordar algunas de lasincertidumbres asociadas con la base de datos.

THE CONCISE INTERNATIONAL CHEMICAL ASSESSMENT DOCUMENT SERIES

Azodicarbonamide (No. 16, 1999)Benzoic acid and sodium benzoate (No. 26, 2000)Benzyl butyl phthalate (No. 17, 1999)Biphenyl (No. 6, 1999)2-Butoxyethanol (No. 10, 1998)Chloral hydrate (No. 25, 2000)Crystalline silica, Quartz (No. 24, 2000)Cumene (No. 18, 1999)1,2-Diaminoethane (No. 15, 1999)3,3'-Dichlorobenzidine (No. 2, 1998)1,2-Dichloroethane (No. 1, 1998)2,2-Dichloro-1,1,1-trifluoroethane (HCFC-123) (No. 23, 2000)Diphenylmethane diisocyanate (MDI) (No. 27, 2000)Ethylenediamine (No. 15, 1999)Ethylene glycol: environmental aspects (No. 22, 2000)2-Furaldehyde (No. 21, 2000)HCFC-123 (No. 23, 2000)Limonene (No. 5, 1998)Manganese and its compounds (No. 12, 1999)Methyl chloride (No. 28, 2000)Methyl methacrylate (No. 4, 1998)Mononitrophenols (No. 20, 2000)Phenylhydrazine (No. 19, 2000)N-Phenyl-1-naphthylamine (No. 9, 1998)1,1,2,2-Tetrachloroethane (No. 3, 1998)1,1,1,2-Tetrafluoroethane (No. 11, 1998)o-Toluidine (No. 7, 1998)Tributyltin oxide (No. 14, 1999)Triglycidyl isocyanurate (No. 8, 1998)Triphenyltin compounds (No. 13, 1999)Vanadium pentoxide and other inorganic vanadium compounds (No. 29, 2001)

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