Toxicology of Industrial Compounds

Toxicology of Industrial Compounds

Toxicology of Industrial Compounds


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<strong>Toxicology</strong> <strong>of</strong> <strong>Industrial</strong> <strong>Compounds</strong>

<strong>Toxicology</strong> <strong>of</strong> <strong>Industrial</strong><br />

<strong>Compounds</strong><br />

Edited by<br />


CIBA-GEIGY Ltd, Basel, Switzerland<br />


Dornach, Switzerland<br />

and<br />


CIBA-GEIGY Ltd, Basel, Switzerland

This edition published in the Taylor & Francis e-Library, 2005.<br />

“To purchase your own copy <strong>of</strong> this or any <strong>of</strong> Taylor & Francis or Routledge’s<br />

collection <strong>of</strong> thousands <strong>of</strong> eBooks please go to www.eBookstore.tandf.co.uk.”<br />

UK Taylor & Francis Ltd, 4 John Street, London WC1N 2ET<br />

USA Taylor & Francis Inc., 1900 Frost Road, Suite 101, Bristol, PA 19007<br />

Copyright © Taylor & Francis Ltd 1995<br />

All rights reserved. No part <strong>of</strong> this publication may be reproduced, stored in a<br />

retrieval system, or transmitted, in any form or by any means, electronic, electro<br />

static, magnetic tape, mechanical, photocopying, recording or otherwise, without<br />

the prior permission <strong>of</strong> the copyright owner.<br />

Library <strong>of</strong> Congress Cataloguing Publication data are available<br />

Cover design by Hybert Design & Type, Maidenhead, Berks.<br />

British Library Cataloguing in Publication Data<br />

A catalogue record for this book is available from the British Library.<br />

ISBN 0-203-97962-1 Master e-book ISBN<br />

ISBN 0-7484-0239-X (Print Edition) (cloth)

Contents<br />

Preface vii<br />

List <strong>of</strong> Contributors ix<br />

PART ONE Bioavailability and metabolic aspects <strong>of</strong> industrial<br />

chemicals<br />

1. Biomonitoring and Absorption <strong>of</strong> <strong>Industrial</strong><br />

Chemicals: the Challenge <strong>of</strong> Organic Solvents<br />

F.A.de Wolff S.Kezic J.G.M.van Engelen<br />

A.C.Monster<br />

2. Toxicokinetics and Biodisposition <strong>of</strong> <strong>Industrial</strong><br />

Chemicals<br />

N.P.E.Vermeulen R.T.H.van Welie B.M.de<br />

Rooij J.N.M.Commandeur<br />

3. Metabolic Activation <strong>of</strong> <strong>Industrial</strong> Chemicals<br />

and Implications for Toxicity<br />

G.J.Mulder<br />

4. Sizing Up the Problem <strong>of</strong> Exposure<br />

Extrapolation: New Directions in Allometric<br />

Scaling<br />

D.B.Campbell<br />

PART TWO Reactive industrial chemicals 59<br />

5. Metabolism <strong>of</strong> Reactive Chemicals<br />

P.J.van Bladeren B.van Ommen<br />

60<br />

6. Methods for the Determination <strong>of</strong> Reactive<br />

<strong>Compounds</strong><br />

P.Sagelsdorff<br />

72<br />

PART THREE Pulmonary toxicology <strong>of</strong> industrial chemicals 90<br />

7. Studies to Assess the Carcinogenic Potential <strong>of</strong><br />

Man-Made Vitreous Fibers<br />

T.W.Hesterberg G.R.Chase R.A.Versen<br />

R.Anderson<br />

91<br />

1<br />

2<br />

12<br />

36<br />


8. Pulmonary Toxicity Studies with Man-Made<br />

Organic Fibres: Preparation and Comparisons<br />

<strong>of</strong> Size-separated Para-aramid with Chrysotile<br />

Asbestos Fibres<br />

D.B.Warheit M.A.Hartsky C.J.Butterick<br />

S.R.Frame<br />

9. Pulmonary Hyperreactivity to <strong>Industrial</strong><br />

Pollutants<br />

J.Pauluhn<br />

10. Mechanisms <strong>of</strong> Pulmonary Sensitization<br />

I.Kimber<br />

11. Occupational Asthma Induced by Chemical<br />

Agents<br />

C.A.C.Pickering<br />

PART FOUR Biomarkers and risk assessment <strong>of</strong> industrial<br />

chemicals<br />

12. Biomarkers and Risk Assessment<br />

K.Hemminki<br />

13. Extrapolation <strong>of</strong> Toxicity Data and Assessment<br />

<strong>of</strong> Risk<br />

N.Fedtke<br />

14. Molecular Approaches to Assess Cancer Risks<br />

A.S.Wright J.P.Aston N.J.van Sittert<br />

W.P.Watson<br />

15. Evaluation <strong>of</strong> Toxicity to the Immune System<br />

H.-W.Vohr<br />

16. New Strategies: the Use <strong>of</strong> Long-term Cultures<br />

<strong>of</strong> Hepatocytes in Toxicity Testing and<br />

Metabolism Studies <strong>of</strong> Chemical Products<br />

Other than Pharmaceuticals<br />

V.Rogiers M.Akrawi S.Coecke<br />

Y.Vandenberghe E.Shephard I.Phillips<br />

A.Vercruysse<br />

117<br />

129<br />

138<br />

149<br />

157<br />

158<br />

167<br />

180<br />

197<br />

207<br />

PART FIVE Mechanisms <strong>of</strong> toxicity <strong>of</strong> industrial chemicals 222<br />

17. Peroxisome Proliferation<br />

B.G.Lake R.J.Price<br />

223<br />


vi<br />

18. Neurotoxicity Testing <strong>of</strong> <strong>Industrial</strong><br />

<strong>Compounds</strong>: in vivo Markers and Mechanisms<br />

<strong>of</strong> Action<br />

K.J.van den Berg J.-B.P.Gramsbergen<br />

E.M.G.Hoogendijk J.H.C.M.Lammers<br />

W.S.Sloot B.M.Kulig<br />

19. Endocrine <strong>Toxicology</strong> <strong>of</strong> the Thyroid for<br />

<strong>Industrial</strong> <strong>Compounds</strong><br />

C.K.Atterwill S.P.Aylward<br />

20. Testing and Evaluation for Reproductive<br />

Toxicity<br />

A.K.Palmer<br />

238<br />

255<br />

280<br />

PART SIX Toxicity <strong>of</strong> selected classes <strong>of</strong> industrial chemicals 300<br />

21. Special Points in the Toxicity Assessment <strong>of</strong><br />

Colorants (Dyes and Pigments)<br />

H.M.Bolt<br />

301<br />

22. <strong>Toxicology</strong> <strong>of</strong> Textile Chemicals<br />

D.Sedlak<br />

309<br />

23. Antioxidants and Light Stabilisers: Toxic<br />

Effects <strong>of</strong> 3,5-Dialkyl-hydroxyphenyl Propionic<br />

Acid Derivatives in the Rat and their Relevance<br />

for Human Safety Evaluation<br />

H.Thomas P.Dollenmeier E.Persohn H.Weideli<br />

F.Waechter<br />

317<br />

24. <strong>Toxicology</strong> <strong>of</strong> Surfactants: Molecular,<br />

Mechanistic and Regulatory Aspects<br />

W.Sterzel<br />

339<br />

PART SEVEN Controversial mechanistic and regulatory issues in<br />

the safety assessment <strong>of</strong> industrial chemicals<br />

25. Low Dose <strong>of</strong> a Genotoxic Carcinogen does not<br />

‘Cause’ Cancer; it Accelerates Spontaneous<br />

Carcinogenesis<br />

W.K.Lutz<br />

26. Controversial Mechanistic and Regulatory<br />

Issues in Safety Assessment <strong>of</strong> <strong>Industrial</strong><br />

Chemicals—an Industry Point <strong>of</strong> View<br />

H.-P.Gelbke<br />

355<br />

356<br />

362<br />

Index 373

Preface<br />

A large number <strong>of</strong> chemical compounds are being constantly introduced<br />

and produced to ease and comfort modern human life. Among those, the<br />

industrial compounds represent that particular fraction <strong>of</strong> chemicals which<br />

are not intended for use in biological systems, but to which humans may be<br />

non-intentionally exposed; at the workplace, by product application or<br />

through the environment.<br />

The International Society for the Study <strong>of</strong> Xenobiotics (ISSX) committed<br />

itself to address, for the first time in the long history <strong>of</strong> industrial<br />

chemicals, the toxicology <strong>of</strong> this class <strong>of</strong> compounds in an intensive<br />

scientific workshop held June 12 through 15, 1994 in Schluchsee,<br />

Germany. This workshop was not only the first such event hosted by ISSX<br />

since its foundation in 1981, but also an extension <strong>of</strong> the society’s scope<br />

beyond its traditionally covered objective to promote studies on xenobiotic<br />

metabolism, disposition and kinetics mainly <strong>of</strong> drugs and agrochemicals.<br />

The large classes <strong>of</strong> pharmaceuticals and agrochemicals had been<br />

deliberately excluded from the scope <strong>of</strong> this workshop, since their terms <strong>of</strong><br />

use generally demand ample registrational toxicity testing that inevitably<br />

leads to a wealth <strong>of</strong> information on, and pr<strong>of</strong>ound toxicological<br />

characterisation <strong>of</strong>, these compounds.<br />

<strong>Industrial</strong> chemicals, instead, which are frequently produced in large<br />

quantities such as pigments, dye-stuffs, plastic materials and additives,<br />

detergents, solvents, etc., to name but a few, are in many cases subjected to<br />

the examination <strong>of</strong> a very basic handling safety only, and may lack any<br />

further toxicity testing. This implies that essentially nothing is known<br />

about their bioavailability, metabolism, excretion and toxicological<br />

properties—unless problems arise. And once toxicity problems come up,<br />

the question arises with them <strong>of</strong> whether or not the available and<br />

traditionally employed methodology is appropriate to approach and solve<br />

them. This, because different from the largely low molecular weight<br />

structures developed for use in biological systems, industrial chemicals are<br />

<strong>of</strong>ten characterised by rather high molecular weight and the incorporation<br />

<strong>of</strong> peculiar structural entities.

viii<br />

Therefore, it was the aim <strong>of</strong> this workshop to contribute to the<br />

investigation <strong>of</strong> industrial chemicals by focussing on the individual<br />

structure, its biological fate, its potential toxicity to mammals and the<br />

molecular mechanisms possibly underlying such adverse effects by<br />

highlighting the use and significance <strong>of</strong> experimental toxicology, with<br />

special emphasis on mechanistic aspects, in the safety assessment <strong>of</strong><br />

industrial compounds as well as to current regulatory and legal<br />

considerations. Topics had been selected to review generally approved facts<br />

and mechanisms, and to particularly address and explore areas <strong>of</strong><br />

investigative and regulatory uncertainty, thereby intending to bring<br />

together the broadly diverse expertise and interests <strong>of</strong> academic<br />

researchers, corporate scientists, experts in safety assessment and<br />

representatives from regulatory authorities.<br />

The following contributions reflect a substantial selection <strong>of</strong> the 27<br />

lectures and six short communications presented during the workshop.<br />

May they succeed in setting a landmark for the due change from the current<br />

era <strong>of</strong> black-box toxicology and largely undifferentiated regulatory<br />

treatment <strong>of</strong> industrial chemicals to the desirable toxicology and safety<br />

assessment by structure in the future.<br />

We gratefully acknowledge the substantial financial support by CIBA-<br />

GEIGY and the RCC Group as well as the financial contributions <strong>of</strong><br />

ADME Bioanalysis, BASF, Henkel, Hüls, Lonza, Schering and Union<br />

Carbide.<br />

Our gratitude is also extended to Mrs Ch.Zehnder for secretarial<br />

assistance and to Taylor & Francis for continuous support, patience and<br />

encouragement to make this publication possible.<br />

H.Thomas<br />

R.Hess<br />


Contributors<br />

May Akrawi<br />

Department <strong>of</strong> Biochemistry and Molecular Biology, University College<br />

London, Gower Street, London WC1E 6BT, UK<br />

Robert Anderson<br />

Schuller MTC, Health, Safety and Environmental Department,<br />

<strong>Toxicology</strong> Group, PO Box 625005, Littleton, CO 80162–5005, USA<br />

J.Paul Aston<br />

Sittingbourne Research Centre, Sittingbourne, Kent ME9 8AG, UK<br />

Christopher K.Atterwill<br />

CellTox Centre, University <strong>of</strong> Hertfordshire, Hatfield Campus, College<br />

Lane, Hatfield AL10 9AB, UK<br />

Samuel P.Aylward<br />

CellTox Centre, University <strong>of</strong> Hertfordshire, Hatfield Campus, College<br />

Lane, Hatfield AL10 9AB, UK<br />

Peter J.van Bladeren<br />

TNO Nutrition and Food Research, PO Box 360, Utrechtseweg 48,<br />

NL-3700 AJ Zeist, The Netherlands<br />

Hermann M.Bolt<br />

Institut für Arbeitsphysiologie, Universität Dortmund, Ardeystrasse 67,<br />

D-44139 Dortmund, Germany<br />

Charles J.Butterick<br />

Texas Technical Health Sciences Centre, Lubbock, TX, USA<br />

D.Bruce Campbell<br />

Servier Research and Development, Fulmer Hall, Windmill Road,<br />

Fulmer, Slough SL3 6HH, UK<br />

Gerald R.Chase<br />

Schuller MTC, Health, Safety and Environmental Department,<br />

<strong>Toxicology</strong> Group, PO Box 625005, Littleton, CO 80162–5005, USA

x<br />

Sandra Coecke<br />

Vrije Universiteit Brussel, Department <strong>of</strong> <strong>Toxicology</strong>, Laarbeeklaan 103<br />

B-1090, Brussels, Belgium<br />

Jan N.M.Commandeur<br />

Division <strong>of</strong> Molecular <strong>Toxicology</strong>, Department <strong>of</strong> Pharmacochemistry,<br />

Vrije Universiteit van Amsterdam, De Boelelaan 1083, H1081 NL-V<br />

Amsterdam, The Netherlands<br />

Peter Dollenmeier<br />

CIBA-GEIGY Ltd., R-1002.2.62, PO Box CH-4002 Basel, Switzerland<br />

Jacqueline G.M.van Engelen<br />

Coronel Laboratory, University <strong>of</strong> Amsterdam, Academic Medical<br />

Centre, Meibergdreef 15, NL-1105 Amsterdam, The Netherlands<br />

Norbert Fedtke<br />

Hüls AG, Bau 2328/PB 12, D-45764 Marl, Germany<br />

Steven R.Frame<br />

DuPont Central Research and Development, Haskell Laboratory, PO<br />

Box 50, Elkton Road, Newark, DE 19714–0050, USA<br />

Heinz-Peter Gelbke<br />

BASF AG, Abt. Toxikologie, D-67056 Ludwigshafen, Germany<br />

Jan-Bert P.Gramsbergen<br />

Department <strong>of</strong> Public Health, Erasmus University, Rotterdam, The<br />

Netherlands<br />

Mark A.Hartsky<br />

DuPont Central Research and Development, Haskell Laboratory, PO<br />

Box 50, Elkton Road, Newark, DE 19714–0050, USA<br />

Kari Hemminki<br />

CNT, Karolinska Institute, Novum, S-141 57 Huddinge, Sweden<br />

Thomas W.Hesterberg<br />

Schuller MTC, Health, Safety and Environmental Department,<br />

<strong>Toxicology</strong> Group, PO Box 625005, Littleton, CO 80162–5005, USA<br />

Elisabeth M.G.Hoogendijk<br />

TNO <strong>Toxicology</strong>, Department <strong>of</strong> Neurotoxicology, PO Box 5815,<br />

NL-2280 HV Rijswijk, The Netherlands<br />

Sanja Keži<br />

Coronel Laboratory, University <strong>of</strong> Amsterdam, Academic Medical<br />

Centre, Meibergdreef 15, NL-1105 Amsterdam, The Netherlands<br />

Ian Kimber<br />

Zeneca Central <strong>Toxicology</strong> Laboratory, Alderley Park, Macclesfield,<br />

Cheshire SK10 4TJ, UK

Beverly M.Kulig<br />

TNO <strong>Toxicology</strong>, Department <strong>of</strong> Neurotoxicology, PO Box 5815,<br />

NL-2280 HV Rijswijk, The Netherlands<br />

Brian G.Lake<br />

BIBRA International, Woodmansterne Road, Carshalton, Surrey, SM5<br />

4DS, UK<br />

Jan H.C.M.Lammers<br />

TNO <strong>Toxicology</strong>, Department <strong>of</strong> Neurotoxicology, PO Box 5815,<br />

NL-2280 HV Rijswijk, The Netherlands<br />

Werner K.Lutz<br />

Universität Würzburg, Institut für Toxikologie, Versbacher Strasse 9,<br />

D-97078 Würzburg, Germany<br />

Aart C.Monster<br />

Coronel Laboratory, University <strong>of</strong> Amsterdam, Academic Medical<br />

Centre, Meibergdreef 15, NL-1105 Amsterdam, The Netherlands<br />

Gerard J.Mulder<br />

Center for Bio-Pharmaceutical Sciences, Sylvius Laboratories, Leiden<br />

University, PO Box 9503, NL-2300 RA Leiden, The Netherlands<br />

Ben van Ommen<br />

TNO Nutrition and Food Research, PO Box 360, Utrechtseweg 48,<br />

NL-3700 AJ Zeist, The Netherlands<br />

Anthony K.Palmer<br />

Huntingdon Research Centre Ltd., PO Box 2, Huntingdon, Cambs,<br />

PE18 6ES UK<br />

Jürgen Pauluhn<br />

BAYER AG, Department <strong>of</strong> <strong>Toxicology</strong>, Institute <strong>of</strong> <strong>Industrial</strong><br />

<strong>Toxicology</strong>, Bldg. 514, D-42096 Wuppertal, Germany<br />

Elke Persohn<br />

CIBA-GEIGY Ltd., Cell Biology Unit, R-1058.2.64, PO Box, CH-4002<br />

Basel, Switzerland<br />

Ian Phillips<br />

Department <strong>of</strong> Biochemistry, Queen Mary and Westfield College,<br />

University <strong>of</strong> London, Mile End Road, London, E1 4NS, UK<br />

C.A.C.Pickering<br />

North West Lung Centre, Wythenshawe Hospital, Southmoor Road,<br />

Manchester M23 9LT, UK<br />

Roger J.Price<br />

BIBRA International, Woodmansterne Road, Carshalton, Surrey SM5<br />

4DS, UK<br />


xii<br />

Vera Rogiers<br />

Vrije Universiteit Brussel, Department <strong>of</strong> <strong>Toxicology</strong>, Laarbeeklaan<br />

103, B-1090 Brussels, Belgium<br />

Ben M.de Rooij<br />

Division <strong>of</strong> Molecular <strong>Toxicology</strong>, Department <strong>of</strong> Pharmacochemistry,<br />

Vrije Universiteit van Amsterdam, De Boelelaan 1083, H1081 NL-V<br />

Amsterdam, The Netherlands<br />

Peter Sagelsdorff<br />

CIBA-GEIGY Ltd., Cell Biology Unit, R-1058.2.52, PO Box, CH-4002<br />

Basel, Switzerland<br />

Dieter Sedlak<br />

Enviro Tex GmbH, Provinostrasse 52, D-86153 Augsburg, Germany<br />

Elizabeth Shephard<br />

Department <strong>of</strong> Biochemistry and Molecular Biology, University College<br />

London, Gower Street, London WC1E 6BT, UK<br />

Nico J.van Sittert<br />

Sittingbourne Research Centre, Sittingbourne, Kent ME9 8AG, UK<br />

Willem S.Sloot<br />

TNO <strong>Toxicology</strong>, Department <strong>of</strong> Neurotoxicology, PO Box 5815,<br />

NL-2280 HV Rijswijk, The Netherlands<br />

Walter Sterzel<br />

Henkel KGaA, TTB-Toxikologie, Geb. Z33, D-40191 Düsseldorf,<br />

Germany<br />

Helmut Thomas<br />

CIBA-GEIGY Ltd., Cell Biology Unit, R-1058.2.46, PO Box, CH-4002<br />

Basel, Switzerland. Current address: Ciba-Pharmaceuticals, Stamford<br />

Lodge, Wilmslow, Cheshire SK9 4LY, UK<br />

Kornelis J.van den Berg<br />

TNO <strong>Toxicology</strong>, Department <strong>of</strong> Neurotoxicology, PO Box 5815,<br />

NL-2280 HV Rijswijk, The Netherlands<br />

Yves Vandenberghe<br />

Vrije Universiteit Brussel, Department <strong>of</strong> <strong>Toxicology</strong>, Laarbeeklaan 103<br />

B-1090 Brussels, Belgium<br />

Antoine Vercruysse<br />

Vrije Universiteit Brussel, Department <strong>of</strong> <strong>Toxicology</strong>, Laarbeeklaan 103<br />

B-1090 Brussels, Belgium<br />

Nico P.E.Vermeulen<br />

Division <strong>of</strong> Molecular <strong>Toxicology</strong>, Department <strong>of</strong> Pharmacochemistry,<br />

Vrije Universiteit Amsterdam, De Boelelaan 1083, H1081 NL-V<br />

Amsterdam, The Netherlands

Richard A.Versen<br />

Schuller MTC, Health, Safety and Environmental Department,<br />

<strong>Toxicology</strong> Group, P.O. Box 625005, Littleton, CO 80162–5005, USA<br />

Hans-Werner Vohr<br />

Bayer AG, Fachbereich Toxikologie, Institut für Toxikologie<br />

Landwirtschaft, Friedrich-Ebert-Strasse 217, D-42096 Wuppertal,<br />

Germany<br />

Felix Waechter<br />

CIBA-GEIGY Ltd, Cell Biology Unit, R-1058.2.68, PO Box, CH-4002<br />

Basel, Switzerland<br />

David B.Wahrheit<br />

DuPont Central Research and Development, Haskell Laboratory, PO<br />

Box 50, Elkton Road, Newark, Delaware 19714–0050, USA<br />

William P.Watson<br />

Sittingbourne Research Centre, Sittingbourne, Kent ME9 8AG, UK<br />

Hansjörg Weideli<br />

CIBA-GEIGY Ltd, R-1002.2.59, PO Box, CH-4002 Basel, Switzerland<br />

Ronald T.H.van Welie<br />

Division <strong>of</strong> Molecular <strong>Toxicology</strong>, Department <strong>of</strong> Pharmacochemistry,<br />

Vrije Universiteit van Amsterdam, De Boelelaan 1083, H1081 NL-V<br />

Amsterdam, The Netherlands<br />

Frederik A.de Wolff<br />

Coronel Laboratory, University <strong>of</strong> Amsterdam, Academic Medical<br />

Centre, Meibergdreef 15, 1105 Amsterdam, The Netherlands<br />

Alan S.Wright<br />

Sittingbourne Research Centre, Sittingbourne, Kent ME9 8AG, UK<br />


PART ONE<br />

Bioavailability and metabolic aspects <strong>of</strong><br />

industrial chemicals

1<br />

Biomonitoring and Absorption <strong>of</strong> <strong>Industrial</strong><br />

Chemicals: the Challenge <strong>of</strong> Organic Solvents<br />



University <strong>of</strong> Amsterdam, Academic Medical Center,<br />

Amsterdam<br />

Introduction<br />

Organic solvents form a very important group <strong>of</strong> industrial chemicals.<br />

They are widely used in a range <strong>of</strong> occupational settings and may exert a<br />

number <strong>of</strong> deleterious effects when subjects are acutely or chronically<br />

exposed. Among the acute effects are skin and mucosal irritation and<br />

general anaesthesia produced by most solvents at high air concentrations.<br />

Examples <strong>of</strong> chronic effects are peripheral neuropathy after long-term<br />

exposure to n-hexane or carbon disulphide, and the organo-psychosyndrome<br />

or ‘solvent dementia’ which may occur after chronic<br />

occupational exposure to a variety <strong>of</strong> volatile organic compounds.<br />

In order to prevent workers from developing solvent-induced<br />

occupational disease, it is essential to set standards for the duration and the<br />

level <strong>of</strong> external exposure. For a scientifically based standard, a clear<br />

understanding is required <strong>of</strong> the relationship between external exposure,<br />

the uptake by the body, the metabolic fate and the internal dose <strong>of</strong> the<br />

substance. The purpose <strong>of</strong> this contribution is to demonstrate the value <strong>of</strong><br />

biokinetic studies in humans to provide a sound scientific basis for<br />

regulatory decisions on occupational standards.<br />

Biological monitoring<br />

In occupational health practice, monitoring is a tool to protect workers<br />

from developing chemically-induced disease. Monitoring in preventive<br />

health care is described as ‘a systemic continuous or repetitive healthrelated<br />

activity, designed to lead if necessary to corrective action’. In<br />

occupational health, a complete monitoring programme consists <strong>of</strong> four<br />

parts: environmental, biological and biological effect monitoring, and<br />

* Also: University Hospital <strong>of</strong> Leiden, Leiden. The Netherlands

F.A.DE WOLFF ET AL. 3<br />

health surveillance. The latter is a major task for the occupational health<br />

physician, but biological monitoring and biological effect monitoring are<br />

fields <strong>of</strong> interest to the occupational toxicologist. In this contribution, only<br />

biological monitoring will be expounded upon.<br />

Biological monitoring (BM) is defined as the ‘measurement and<br />

assessment <strong>of</strong> workplace agents or their metabolites either in tissues,<br />

secreta, excreta or any combination <strong>of</strong> these to evaluate exposure and<br />

health risk compared to an appropriate reference’ (Zielhuis & Henderson,<br />

1986). This means that a biological monitoring programme is not limited<br />

to the assay <strong>of</strong> xenobiotics in biological samples. As in clinical laboratory<br />

medicine, the pre-analytical phase <strong>of</strong> the process is very important, and<br />

even more so the post-analytical phase <strong>of</strong> the laboratory analysis, which<br />

means the interpretation <strong>of</strong> the analytical data in biomedical terms. The<br />

ultimate goal <strong>of</strong> biological monitoring is the evaluation <strong>of</strong> the health risk <strong>of</strong><br />

workers by estimation <strong>of</strong> the internal dose <strong>of</strong> a chemical. This is not limited<br />

to measurement <strong>of</strong> the quantity <strong>of</strong> the substance absorbed by the body, but<br />

may also include the assay <strong>of</strong> metabolites <strong>of</strong> toxicological interest, if<br />

possible in or near a critical organ (Monster & van Hemmen, 1988).<br />

This implies that the absorption, metabolism and elimination <strong>of</strong> a<br />

substance in man should be known before a biological monitoring<br />

programme can be performed in practice. Animal experiments are <strong>of</strong><br />

limited value; volunteer studies in order to determine pulmonary and<br />

dermal uptake <strong>of</strong> organic solvents provide more relevant data for this<br />

purpose.<br />

Owing to the existence <strong>of</strong> very sensitive analytical methods it is possible<br />

to study the kinetics and metabolism <strong>of</strong> solvents in volunteers who are<br />

experimentally exposed to levels at or far below the <strong>of</strong>ficial threshold limit<br />

values, so that any health risk for the volunteers can almost totally be<br />

excluded.<br />

As with biological monitoring <strong>of</strong> most other substances, in the case <strong>of</strong><br />

organic solvents the compound itself and/or its metabolite in blood or urine<br />

can be measured. Studies with volatile, rather lipophilic, substances have an<br />

additional advantage, namely that the solvent can also be measured in<br />

expired air. Analytically this has the advantage <strong>of</strong> an extremely clean<br />

matrix in comparison with body fluids, whereas biologically, air samples<br />

provide us with information on the blood concentration <strong>of</strong> a volatile<br />

compound. Moreover, collection <strong>of</strong> expired air is non-invasive and large<br />

volumes are readily available (Droz & Guillemin, 1986).<br />

An example <strong>of</strong> a study on solvents in volunteers is the one carried out in<br />

our laboratory on the biokinetics <strong>of</strong> n-hexane and its neurotoxic metabolite<br />

2,5-hexanedione (Van Engelen et al., in preparation). Volunteers are<br />

exposed during 15 min to 60 ppm hexane by inhalation. The minute volume<br />

and the respiratory rate are measured and blood and exhaled air sampled<br />

frequently for determination <strong>of</strong> 2,5-hexanedione and n-hexane,


respectively. Each volunteer is exposed twice in succession on one test day<br />

in order to get an impression <strong>of</strong> the within-day intra-individual variation.<br />

Venous blood is sampled through a catheter, and alveolar air is collected<br />

after holding breath for 30 s (to achieve equilibrium between pulmonary<br />

blood and air) by exhaling through a glass tube which is stoppered<br />

immediately. These tubes contain 70 ml alveolar air and the total volume is<br />

analyzed for n-hexane by using a purge-and-trap system. 2,5-Hexanedione<br />

in serum is measured by using electron capture detection after<br />

derivatization, with a detection limit <strong>of</strong> 30 micro-mol l −1 (Keži and<br />

Monster, 1991). During exposure the concentration <strong>of</strong> n-hexane in<br />

alveolar air increases very rapidly and decreases after discontinuation <strong>of</strong><br />

exposure. The half-life time <strong>of</strong> exhalatory elimination after the distribution<br />

phase is in the order <strong>of</strong> 30 min.<br />

2,5-Hexanedione becomes detectable in blood as fast as 2–3 min after<br />

commencement <strong>of</strong> n-hexane exposure. After discontinuation <strong>of</strong> dosing the<br />

metabolite concentration continues to increase for another 3 min, to<br />

disappear from the plasma with a half-life <strong>of</strong> approximately 1.5 h. The<br />

second exposure period on the same day shows very reproducible n-hexane<br />

and 2,5-hexanedione curves in the same individual. Between individuals<br />

there is considerable variation in kinetics and metabolism, and this issue is<br />

being studied in detail at present.<br />

Before a biological monitoring programme can be designed, a detailed<br />

biokinetic study like this one, <strong>of</strong> every solvent being used in industry, has to<br />

be performed. Without kinetic data it is impossible to choose for instance<br />

the correct matrix, the compound to be measured, or the sampling<br />

frequency and time. In addition, these data are necessary to establish a<br />

relationship between ambient air concentrations <strong>of</strong> a chemical (external<br />

exposure), and the biological parameters used to estimate a health risk.<br />

Absorption<br />

The primary association <strong>of</strong> the pharmacologist or general toxicologist,<br />

when reading or hearing the term ‘absorption’, is with ‘intestinal’. For<br />

drugs, gastrointestinal uptake is indeed the most common route to enter<br />

the body. In case <strong>of</strong> occupational exposure, however, intestinal absorption<br />

is <strong>of</strong> minor importance. The occupational toxicologist is, therefore, more<br />

inclined to pay attention to entry routes other than the intestine, the most<br />

important being pulmonary and dermal uptake.<br />

Pulmonary uptake<br />

There are a number <strong>of</strong> parameters which affect the pulmonary uptake <strong>of</strong><br />

organic solvents. In the first place, the physical chemistry <strong>of</strong> the compound<br />

is <strong>of</strong> importance. Both the blood-to-gas and the tissue-to-blood partition

F.A.DE WOLFF ET AL. 5<br />

Figure 1.1 The mean minute volume (1 min −1 ) and the percentage <strong>of</strong> the minute<br />

volume cleared from solvent (shaded area) during exposure to styrene (left) and 1,1,<br />

1-trichloroethane (right) at increasing degree <strong>of</strong> workload.<br />

coefficients determine the absorption through the alveolar membrane and<br />

the distribution over the body. Furthermore, exercise is an important<br />

physiological determinant. With increasing exercise, ventilation increases<br />

and, therefore, also the availability <strong>of</strong> the vapour to the lung per unit <strong>of</strong><br />

time. In addition, cardiac output increases during exercise, and this may<br />

affect absorption, distribution and metabolism through enhanced blood<br />

flow.<br />

Finally, the elimination <strong>of</strong> a solvent which occurs during exposure may<br />

significantly affect the uptake rate. The percentage <strong>of</strong> the vapour not<br />

retained by the body but exhaled again is dependent on, again,<br />

physicochemical factors such as solubility, but also on the rate <strong>of</strong><br />

metabolism (Fiserova-Bergerova, 1985).<br />

In order to demonstrate the different factors which may affect<br />

pulmonary absorption <strong>of</strong> vapours we have constructed Figure 1.1, based<br />

on earlier work <strong>of</strong> Astrand et al. (Astrand, 1975). In their studies,<br />

volunteers were exposed to different vapours such as styrene or 1,1,1trichloroethane<br />

at increasing degrees <strong>of</strong> workload during 2 h.<br />

The first 30 min they were exposed at rest, and then the workload was<br />

increased every 30 min with 50 W. The minute volume, here referred to as<br />

‘supply’, was measured and expressed in 1 min −1 , and the exhaled solvent<br />

concentration was also measured at regular intervals. The shaded area <strong>of</strong><br />

the vertical bars in Figure 1.1 indicate the percentage <strong>of</strong> minute volume<br />

cleared from the solvent, averaged over the observation period. This is<br />

considered to be a measure for pulmonary uptake.


During continuous exposure to a constant concentration and at<br />

increasing exercise the uptake <strong>of</strong> styrene remains constant, expressed in<br />

terms <strong>of</strong> percentage <strong>of</strong> the minute volume cleared. Apparently, the body is<br />

not easily saturated with styrene. The picture for 1,1,1-trichloroethane is<br />

completely different. Although the minute volume at each level <strong>of</strong><br />

workload is comparable with that <strong>of</strong> the styrene experiment, it is clear that<br />

the retention <strong>of</strong> 1,1,1-trichloro-ethane is much lower. Apparently, the body<br />

becomes rapidly saturated with 1,1,1-trichloroethane. The reasons for the<br />

difference in pulmonary uptake between these two solvents are evident.<br />

Styrene is highly soluble in blood and it is extensively metabolized to<br />

mandelic acid and phenyl glyoxylic acid. The retention in the body remains<br />

the same, and therefore the uptake increases proportionally with the<br />

minute volume.<br />

In contrast to styrene, 1,1,1-trichloroethane has only a limited solubility<br />

in blood, and it is hardly metabolized. This means that during exposure the<br />

body becomes rapidly saturated with the substance, and that an increase in<br />

minute volume by increasing workload results in a lower retention, and<br />

hardly in higher uptake. Differences in kinetic behaviour, as demonstrated<br />

for styrene and 1,1,1-trichloroethane, are important for the design <strong>of</strong> a<br />

biological monitoring programme.<br />

Dermal uptake<br />

Absorption <strong>of</strong> solvents through the skin may be affected by a number <strong>of</strong><br />

factors. Many organic solvents are able to penetrate the skin and thus enter<br />

the body. This is a rather well-known fact which can be prevented in<br />

industrial practice by use <strong>of</strong> protective clothing. It is, however, less<br />

common knowledge that solvents in the vapour phase may also penetrate<br />

the skin. In case <strong>of</strong> skin exposure to liquids usually a small surface is<br />

exposed, whereas in case <strong>of</strong> vapour the whole body surface <strong>of</strong> about 2 m 2<br />

may be exposed. This means that under certain conditions skin absorption<br />

<strong>of</strong> vapour may significantly contribute to the amount absorbed by<br />

inhalation.<br />

Other parameters which may affect skin absorption are the temperature,<br />

and the ability <strong>of</strong> some solvents to increase their own absorption by<br />

causing skin hyperaemia through irritation. To demonstrate these factors,<br />

some preliminary results are shown <strong>of</strong> a volunteer study on skin<br />

penetration <strong>of</strong> solvents in the liquid and vapour phases (Keži et al., in<br />

preparation).<br />

The experimental conditions are as follows. The volunteer is seated in a<br />

clear-air cabin in order to avoid additional inhalatory exposure to vapour<br />

in the experimental room. The arm is the only part <strong>of</strong> the body outside the<br />

cabin. In case <strong>of</strong> exposure to liquid on the skin, the solvent is put in a<br />

chamber which is pressed on to the skin during the exposure period, which

F.A.DE WOLFF ET AL. 7<br />

is usually no longer than a few minutes. The exposed area is usually in the<br />

order <strong>of</strong> 20 cm 2 .<br />

In the case <strong>of</strong> dermal exposure to vapour, the volunteer places the lower<br />

arm into a piece <strong>of</strong> drainage pipe through which the vapour is led with<br />

controlled flow and concentration in air. Uptake <strong>of</strong> liquid or vapour is<br />

measured in both cases by determination <strong>of</strong> the solvent in expired air, by<br />

the sampling method described earlier.<br />

Figure 1.2 shows the dermal uptake and elimination <strong>of</strong> two different<br />

liquids in one volunteer. A surface <strong>of</strong> 27 cm 2 was exposed during 3 min to<br />

pure 1,1.1-trichloroethane and to tetrachloroethene. It is clear that 1,1,1trichloroethane<br />

is absorbed through the skin much faster and to a much<br />

greater extent than tetrachloroethene, at least in exposure to the liquids.<br />

However, when the skin is exposed to the same solvents in the vapour<br />

phase the picture becomes totally different. Here the lower arm, which has<br />

a surface <strong>of</strong> about 500 cm 2 , was exposed during 15 min to solvent<br />

concentrations <strong>of</strong> approximately 500 µmol 1 −1 air (Figure 1.3).<br />

In the case <strong>of</strong> vapour exposure no difference in absorption kinetics is<br />

observed, and only a small difference in expired air concentration is seen.<br />

The reason for the discrepancy between vapour exposure is that 1,1,1trichloro-ethane<br />

causes skin irritation as the liquid, but not in the vapour<br />

phase. Irritation leads to hyperaemia and, hence, increased absorption.<br />

As it is known that dermal exposure to vapour may lead to detectable<br />

absorption, the contribution <strong>of</strong> vapour uptake <strong>of</strong> the skin in comparison to<br />

inhalatory absorption should be evaluated. This was done with<br />

trichloroethene as an example (Figure 1.4). Both curves were obtained in<br />

the same volunteer. The dermal exposure was performed first, followed by<br />

the inhalatory test after a wash out period <strong>of</strong> 2 weeks. The exposure period<br />

was 15 min, and the inhalatory concentration was 4.1 µmol l −1 . Dermal<br />

exposure <strong>of</strong> the lower arm took place at 1.4 mmol l −1 .<br />

It appears that uptake from the lungs occurs much faster than via the<br />

skin. This is conceivable because the stratum corneum is a stronger barrier<br />

than the alveolar epithelium, and causes a shift to the right <strong>of</strong> the t max. It<br />

can also be seen that inhalatory exposure leads to a much higher expired<br />

air concentration than dermal exposure. But in this respect we should<br />

realize that only a small part <strong>of</strong> the skin was exposed, namely about 500<br />

cm 2 . In fact the result should be extrapolated to the total surface <strong>of</strong> the<br />

human skin, which is about 2 m 2 . These results indicate that dermal<br />

exposure to solvent vapour should not be neglected when the safety <strong>of</strong> the<br />

industrial environment is evaluated. This is <strong>of</strong> special importance when<br />

ambient air concentrations are high, and workers are protected with<br />

protective masks but not with gloves. Another example in which skin<br />

absorption may be high in comparison with inhalation are those solvents<br />

which are readily absorbed by the skin, such as 2-butoxyethanol (Johanson<br />

and Boman, 1991).


Figure 1.2 Elimination <strong>of</strong> 1,1,1-trichloroethane and tetrachloroethene by expired<br />

air after dermal exposure to the liquid <strong>of</strong> 27 cm 2 fore-arm skin during 3 min. 1,1,1trichloroethane<br />

liquid irritates the skin.<br />

Figure 1.3 Elimination <strong>of</strong> 1,1,1-trichloroethane and tetrachloroethene by expired<br />

air after dermal exposure to the vapour <strong>of</strong> 500 cm 2 lower-arm skin during 15 min<br />

to 500 µmol l −1 air.<br />

The temperature <strong>of</strong> the solvent is another factor that may have an<br />

influence on uptake through the skin. Figure 1.5 shows the results <strong>of</strong><br />

dermal exposure to liquid tetrachloroethene and n-hexane at two different<br />

temperatures in one volunteer. Exposure time here was only 1 min, and<br />

absorption and elimination were measured by analysis <strong>of</strong> the vapours in<br />

expired air.

At the low temperature <strong>of</strong> the liquid (15°C), the uptake <strong>of</strong><br />

tetrachloroethene is negligible when compared with a normal skin<br />

temperature <strong>of</strong> 33°C. In case <strong>of</strong> n-hexane, under comparable circumstances<br />

and in the same volunteer, the effect <strong>of</strong> temperature is much less<br />

pronounced. Apparently, the physicochemical properties <strong>of</strong> the solvent are<br />

an additional determining factor. The mechanism on which the difference<br />

between tetrachloroethene and n-hexane is based is the subject <strong>of</strong> further<br />

study.<br />

Conclusions<br />

F.A.DE WOLFF ET AL. 9<br />

Figure 1.4 Elimination <strong>of</strong> trichloroethene by expired air during and after inhalatory<br />

exposure to 4.1 µmol l −1 trichloroethene during 15 min, and after dermal vapour<br />

exposure during 15 min <strong>of</strong> the lower-arm skin (500 cm 2 to 1.4 mmol l −l air).<br />

In occupational health practice, the major absorption routes for organic<br />

solvents are not ingestion, but inhalation and skin penetration, the latter<br />

both as liquid and as vapour. The physical chemistry <strong>of</strong> the compound,<br />

exercise, and the elimination rate may affect pulmonary uptake. Factors<br />

affecting dermal uptake are the ability <strong>of</strong> the solvent to penetrate the skin as<br />

liquid or vapour, the temperature <strong>of</strong> the liquid, and the irritability <strong>of</strong> the<br />

chemical to the skin.<br />

Before a biological monitoring programme for solvent exposure can be<br />

set up, the kinetics and metabolism <strong>of</strong> the various solvents in man should<br />

be known. Owing to the availability <strong>of</strong> sensitive analytical methods it is<br />

usually possible to perform volunteer studies at safe exposure levels.<br />

Measurement <strong>of</strong> solvents in expired air and <strong>of</strong> their metabolites in body<br />

fluids is <strong>of</strong> the utmost importance to estimate the internal dose <strong>of</strong> the<br />

solvents and health risk to which man can be exposed in the work and<br />

general environment.


Figure 1.5 Elimination <strong>of</strong> tetrachloroethene and n-hexane by expired air after<br />

dermal exposure during 1 min to liquid at 15°C and 33°C<br />

References<br />

ǺSTRAND, I., 1975, Uptake <strong>of</strong> solvents in the blood and tissues in man. A review,<br />

Scand J Work Environ Health, 1, 199–218.<br />

DROZ, P.O. and GUILLEMIN, M.P., 1986, Occupational exposure monitoring<br />

using breath analysis, J Occup Med, 28, 593–602.

F.A.DE WOLFF ET AL. 11<br />

FISEROVA-BERGEROVA, V., 1985, Toxicokinetics <strong>of</strong> organic solvents, Scand J<br />

Work Environ Health, 11, suppl. 1, 7–21.<br />

JOHANSON, G. and BOMAN, A., 1991, Percutaneous absorption <strong>of</strong> 2butoxyethanol<br />

vapour in human subjects, Br J Ind Med, 48, 788–92.<br />

KEŽI , S. and MONSTER, A.C., 1991, Determination <strong>of</strong> 2,5-hexanedione in urine<br />

and serum by gaschromatography after derivatization with O-<br />

(pentafluorobenzyl)-hydroxylamine and solid-phase extraction, J Chromatogr,<br />

563, 199–204.<br />

MONSTER, A.C. and VAN HEMMEN, J.J., 1988, Screening models in<br />

occupational health practice <strong>of</strong> assessment <strong>of</strong> individual exposure and health<br />

risk by means <strong>of</strong> biological monitoring in exposure to solvents, In Notten,<br />

W.R.F., Herber, R.F. M., Hunter, W.J. et al. (Eds) Health Surveillance <strong>of</strong><br />

lndividual Workers Exposed to Chemical Agents, pp. 47–53, Berlin: Springer.<br />

ZIELHUIS, R.L. and HENDERSON, P.Th., 1986, Definitions <strong>of</strong> monitoring<br />

activities and their relevance for the practice <strong>of</strong> occupational health, Int Arch<br />

Occup Environ Health, 57, 249–57.

2<br />

Toxicokinetics and Biodisposition <strong>of</strong> <strong>Industrial</strong><br />

Chemicals<br />



Vrije Universiteit, Amsterdam<br />

Introduction<br />

In our industrialized world with increasing numbers <strong>of</strong> body foreign<br />

chemicals (xenobiotics) including drugs, food additives, pesticides,<br />

industrial chemicals and environmental pollutants, public concern about<br />

possible adverse (health) effects is growing. In 1989, for example, actual<br />

environmental topics in the Netherlands were photochemical summersmog<br />

and the presence <strong>of</strong> dioxines in milk <strong>of</strong> cows feeding in the neighbourhood<br />

<strong>of</strong> household refuse combustion furnaces and cable stills (CCRX, 1989). In<br />

this regard, most attention is paid to exposure to potentially mutagenic and<br />

carcinogenic xenobiotic chemicals. Apart from environmental exposure,<br />

especially at the workplace man may be exposed to elevated levels <strong>of</strong><br />

mixtures <strong>of</strong> known or unknown chemicals. Two centuries ago, cancer <strong>of</strong> the<br />

scrotum and testicles in chimney-sweepers was the first recognized<br />

occupational cancer (Pott, 1795). Since then numerous other hazardous<br />

occupational activities have been traced (Farmer et al., 1987).<br />

Nowadays, toxicologists are more and more focussed on the in vivo and<br />

in vitro bioactivation and bioinactivation mechanisms <strong>of</strong> chemicals. In the<br />

development <strong>of</strong> toxicity different stages are generally being distinguished:<br />

(1) toxicokinetics (absorption, distribution and elimination), (2)<br />

biotransformation, resulting in activation or inactivation <strong>of</strong> the chemicals,<br />

(3) reversible or irreversible interactions with cellular or tissue<br />

components, (4) protection and repair mechanisms and (5) nature and<br />

extent <strong>of</strong> the toxic effect for the organism (Vermeulen et al., 1990).<br />

Knowledge <strong>of</strong> for example species, dose, route <strong>of</strong> absorption, time <strong>of</strong><br />

exposure, tissue and organ selective interactions with (critical) cellular<br />

macro-molecules contributes to the understanding <strong>of</strong> molecular mechanisms<br />

<strong>of</strong> toxicity. Molecular mechanisms are useful in the prediction and<br />

prevention <strong>of</strong> chemically induced toxicities and they may play an<br />

important role in for example risk assessment and in the development <strong>of</strong><br />

safer chemicals (Vermeulen et al., 1990).

In this chapter, first the basic toxicokinetic concepts concerning the dis<br />

tribution, elimination and biotransformation <strong>of</strong> xenobiotics will be<br />

summarized. Subsequently, the relevance <strong>of</strong> these concepts will be<br />

illustrated and evaluated with the aid <strong>of</strong> a number <strong>of</strong> toxicokinetic studies<br />

in animals and humans concerning the nematocide 1,3-dichloropropene,<br />

the fungicide etridiazol, the chemical monomer 1,3-butadiene and the<br />

industrial solvent, 1,1,2-tri-chloroethylene. Apart from interspecies<br />

differences in the toxicokinetics, special attention will be given to<br />

interindividual differences in the toxicokinetics, among other things, as a<br />

result <strong>of</strong> genetically determined deficiencies in biotransformation enzymes<br />

as well as to its importance for the risk assessment <strong>of</strong> human exposure to<br />

industrial chemicals.<br />

Disposition <strong>of</strong> xenobiotics<br />


The overall fate <strong>of</strong> xenobiotics in an organism is determined by various<br />

toxicokinetic processes notably the route <strong>of</strong> administration, absorption,<br />

distribution and elimination. Chemicals may enter the body via various<br />

routes. Main routes are the lung, skin and gastrointestinal tract. The<br />

intraperitoneal, intramuscular, intravenous and subcutaneous routes are<br />

largely confined to experimental toxicological and therapeutic agents.<br />

Following absorption, xenobiotics enter the systemic or portal blood<br />

circulation. Distribution <strong>of</strong> chemicals in blood, organs and tissues usually<br />

occurs rapidly. The final plasma concentration depends on the ability <strong>of</strong><br />

the chemicals to pass cell membranes and on their affinity to various<br />

macromolecular proteins and tissues. Distribution to the kidney may result<br />

in direct excretion <strong>of</strong> the unchanged parent chemical. The physicochemical<br />

characteristics, such as lipophilicity and binding to plasma proteins, play an<br />

important role in the ultimate fate <strong>of</strong> a chemical in the body. The<br />

disposition <strong>of</strong> xenobiotics in the body is shown schematically in Figure 2.1.<br />

Its schematic relationship with biological/ toxicological effects is shown in<br />

Figure 2.2.<br />

Biotransformation plays an important role in the disposition <strong>of</strong><br />

xenobiotics in vivo. The liver is quantitatively the most important organ in<br />

the process <strong>of</strong> biotransformation. It receives a relative high bloodflow<br />

directly from the gastrointestinal tract via the portal vein, sometimes giving<br />

rise to the so-called hepatic ‘first-pass effect’ due to the presence <strong>of</strong> high<br />

concentrations <strong>of</strong> phase I and phase II metabolizing enzymes.<br />

Other important organs in biotransformation are the lungs, kidneys and<br />

the intestine. The primary object <strong>of</strong> biotransformation generally is to<br />

increase the hydrophilicity <strong>of</strong> chemicals, thus facilitating excretion by the<br />

kidneys in the urine or by the liver in the bile. Phase I reactions involve<br />

oxidation, reduction and hydrolysis reactions and phase II reactions<br />

conjugation or synthetic reactions. Phase I metabolic reactions generally


Figure 2.1 Schematic representation <strong>of</strong> the fate <strong>of</strong> xenobiotics in the body according<br />

to their physico-chemical properties. Phase I and phase II represent the<br />

biotransformation processes. Adapted from Ariens and Simonis (1980).<br />

convert xenobiotic chemicals to more hydrophilic derivatives by<br />

introducing functional groups such as hydroxyl, sulphydryl and amino- or<br />

carboxylic acid groups. Phase II reactions are conjugation reactions in<br />

which the parent compounds or phase I derived metabolites are covalently<br />

bound to for example glucuronic acid, sulphate or glutathione.<br />

The group <strong>of</strong> cytochrome P-450 isoenzymes is the most important enzyme<br />

system in the catalysis <strong>of</strong> phase I reactions. The microsomal cytochrome<br />

P-450 system consists <strong>of</strong> various cytochrome P-450 isoenzymes and<br />

NADPH-cytochrome P450 reductase. It is involved in different metabolic<br />

reactions. At least three main types <strong>of</strong> activities can be distinguished,<br />

namely monooxygenase activity, oxidase activity and reductive activity<br />

(Guengerich 1994; Koymans et al., 1993). Glucuronic acid conjugation,<br />

catalyzed by UDP-glucuronyltransferases, represents one <strong>of</strong> the major<br />

phase II conjugation reactions in the conversion <strong>of</strong> exogenous and

endogenous chemicals. In mammals, another important conjugation<br />

reaction <strong>of</strong> hydroxyl groups is sulfatation, catalyzed by sulfotransferases<br />

(Sipes and Gandolfi, 1986). The group <strong>of</strong> glutathione S-transferase (GST)<br />

isoenzymes also represents an important phase II enzyme system. GST<br />

isoenzymes consist <strong>of</strong> two subunits on which the nomenclature is based<br />

(Warholm et al., 1986). The most important activity <strong>of</strong> GSTs is the<br />

catalysis <strong>of</strong> the conjugation <strong>of</strong> electrophilic, hydrophobic chemicals with<br />

the tripeptide glutathione (GSH). In general, GSH conjugation ultimately<br />

leads to the urinary excretion <strong>of</strong> mercapturic acids (N-acetyl-L-cysteine Sconjugates)<br />

(Vermeulen, 1989; Van Welie et al., 1992).<br />

Toxicokinetic principles<br />

General principles<br />


Figure 2.2 Disposition and biological effects <strong>of</strong> xenobiotics subdivided into three<br />

phases.<br />

The time course for the absorption, distribution, metabolism and<br />

elimination <strong>of</strong> a toxic substance is the subject <strong>of</strong> toxicokinetics. Implicit in<br />

any toxicokinetic description is the assumption that the response <strong>of</strong> target<br />

tissues or organs can be related to concentration pr<strong>of</strong>iles <strong>of</strong> the active form<br />

<strong>of</strong> the substance in that tissue or organ. Furthermore, it is <strong>of</strong>ten assumed<br />

that blood or plasma concentrations in one way or the other will reflect<br />

target tissue or organ concentrations and by inference the toxic effects.<br />

Under normal conditions one is generally dealing with first-order or linear<br />

kinetics, meaning that the amount <strong>of</strong> compound absorbed or eliminated<br />

(dQ) per unit <strong>of</strong> time (dt) is proportional to the total amount <strong>of</strong> compound<br />

present in the body. Zeroorder or non-linear kinetics may be valid as a<br />

consequence <strong>of</strong> various causes, e.g. saturation <strong>of</strong> binding <strong>of</strong> the toxic<br />

substance to plasma proteins or tissue components, or, more frequently


Table 2.1 Frequently used toxicokinetic parameters and their formulas<br />

occurring, saturation <strong>of</strong> biotransformation enzyme systems. For the<br />

(mathematical) description <strong>of</strong> the toxicokinetics <strong>of</strong> substances, there exist<br />

at least two approaches at the moment: the traditional compartment<br />

pharmaco(toxico-)kinetic approach, in which the body is divided into one<br />

or more compartments, which do not necessarily correspond to<br />

physiological or anatomical units, and the physiologically-based pharmaco<br />

(toxico-)kinetic approach (PBPK or PBTK), in which organs, tissues and<br />

blood flow are taken into consideration. In Table 2.1 a summary <strong>of</strong> the most<br />

important and most frequently used traditional toxicokinetic parameters is<br />

shown. The value <strong>of</strong> some <strong>of</strong> these parameters is illustrated below, with the<br />

examples <strong>of</strong> 1,3-dichloropropene and etridiazol. The PBPK/PBTK approach<br />

is illustrated with the example <strong>of</strong> 1,3-butadiene.<br />

Principles <strong>of</strong> urinary excretion<br />

Of special interest in relation to this contribution also is the urinary<br />

excretion <strong>of</strong> xenobiotics and their metabolites by the kidneys. Two basic


processes, namely glomerular filtration and tubular secretion are used by<br />

the kidneys to remove chemicals from the bloodstream into the urine<br />

(Hook and Hewitt 1986). The kidneys are highly vulnerable to potential<br />

toxicants not only because they receive a high bloodflow (25% <strong>of</strong> the<br />

cardiac output), but also because they have the intrinsic ability to<br />

concentrate compounds. Recently, it has also become clear that xenobiotics<br />

may become nephrotoxic in the kidney itself due to bioactivation processes<br />

in combination with insufficient protection mechanisms (Commandeur and<br />

Vermeulen, 1991).<br />

The elimination <strong>of</strong> chemicals by the kidney is generally governed by firstorder<br />

processes. During first-order excretion kinetics the urinary<br />

elimination rate <strong>of</strong> a chemical is directly proportional to the plasma<br />

concentration. This means that the higher the plasma concentration the<br />

more <strong>of</strong> the chemical will be excreted in urine per unit <strong>of</strong> time. The urinary<br />

elimination rate (dQ/dt) can be calculated from a semi-logarithmic plot <strong>of</strong><br />

the urinary elimination rate versus the time <strong>of</strong> the intermittently collected<br />

urine samples (dQ/dt (mg h −l )=volume (1)×concentration (mg 1 −1 )/time (h))<br />

(Figure 2.3A).<br />

From the slope <strong>of</strong> the semi-logarithmic plasma concentration or urinary<br />

excretion rate versus time curve, the elimination rate constant (k el) and the<br />

urinary half-life <strong>of</strong> elimination (t 1/2) can be calculated. The half-life <strong>of</strong><br />

elimination is the time required to decrease the plasma concentration or the<br />

urinary elimination rate by one-half. The volume <strong>of</strong> distribution <strong>of</strong> the<br />

chemical normally can not be calculated from the urinary excretion data.<br />

Because the amount <strong>of</strong> chemical excreted in urine per unit <strong>of</strong> time (dQ/dt)<br />

is proportional to the plasma concentration (C p), the t 1/2 derived from the<br />

urinary elimination rate constant is identical to the t 1/2 <strong>of</strong> the chemical in<br />

plasma. It is evident that under these conditions the urinary excretion rate<br />

curve has the same shape as the plasma concentration curve (Figure 2.3B).<br />

In practice, the concentration <strong>of</strong> a chemical in urine (mg l −1 ) can be<br />

determined and multiplied by the volume (1) <strong>of</strong> the urine sample in order<br />

to calcu late the amount (mg) <strong>of</strong> chemical excreted over a period <strong>of</strong> time. In<br />

a semi-logarithmic plot the amount <strong>of</strong> chemical excreted is plotted against<br />

the midpoint <strong>of</strong> the interval <strong>of</strong> collection (Figure 2.3B). The accuracy <strong>of</strong> the<br />

method strongly depends on the way and the number <strong>of</strong> urine samples<br />

collected. As a rule <strong>of</strong> thumb, urine samples have to be collected during at<br />

least four half-lives <strong>of</strong> elimination. The complete cumulative urinary<br />

excretion <strong>of</strong> a chemical can be calculated as the area under the urinary<br />

excretion rate versus time curve including extrapolation time to infinity.<br />

Occupational exposure to chemicals frequently occurs 5 days a week, 8 h<br />

a day, with an exposure free period <strong>of</strong> 16 h. Intermittent exposure to a<br />

chemical may lead to different accumulation situations in the body<br />

depending on the periods between exposure in relation to t 1/2 (Table 2.1).<br />

No accumulation will occur when the intervals between the exposure


Figure 2.3 Schematic representation <strong>of</strong> first order kinetics <strong>of</strong> (A) the plasma<br />

concentration (C p) <strong>of</strong> a chemical versus the urinary elimination rate (dQ/dt), (B) the<br />

relation between the elimination rate in plasma and urine and (C) the cumulative<br />

excretion ( (%)) versus time. In (B): slope=–k el/2.303 and t 1/2=0.693/k el.<br />

Figure 2.4 Urinary excretion <strong>of</strong> a hypothetical metabolite during 3 days <strong>of</strong><br />

intermittent exposure: t 1/2

urinary concentration at certain time points the net total cumulative<br />

excretion <strong>of</strong> day 2 can be calculated.<br />

Monitoring in occupational toxicology<br />


In occupational toxicology generally four monitoring approaches are<br />

distinguished, namely: environmental monitoring (EM), biological<br />

monitoring (BM), biological effect monitoring (BEM) and health<br />

surveillance (HS) (Figure 2.5). EM and BM are concerned with the<br />

measurement and assessment <strong>of</strong> ambient exposure and health risk<br />

compared to appropriate references. EM determines xenobiotics at the<br />

workplace, BM determines xenobiotics or their metabolites in tissues or<br />

secreta. BEM is concerned with the measurement and assessment <strong>of</strong> early,<br />

non-adverse, biological alterations in exposed workers to evaluate<br />

exposure and/or health risk compared to appropriate references. HS is<br />

concerned with periodic medico-physiological examination <strong>of</strong> exposed<br />

workers with the objective <strong>of</strong> protecting and preventing occupationally<br />

related diseases (Zielhuis and Henderson, 1986).<br />

EM was shown to be <strong>of</strong> limited value for assessing the internal dose <strong>of</strong> a<br />

chemical by not taking into account for example toxicokinetic and<br />

toxicodynamic processes determining the ultimate fate <strong>of</strong> xenobiotics in the<br />

body. To a certain extent, BM appeared to overcome the problems<br />

inherently related to EM. BM assesses the overall exposure to xenobiotics<br />

that are present at the workplace through measurement <strong>of</strong> the appropriate<br />

determinant(s) in biological specimens collected from the worker at specific<br />

timepoints (ACGIH, 1990).<br />

Ideally, not only the relation between exposure and effect is known, but<br />

also the toxicokinetic and toxicodynamic interactions linking these two. If<br />

these processes are elucidated, quantitative knowledge <strong>of</strong> a determinant <strong>of</strong><br />

one <strong>of</strong> the different monitoring methods allows an assessment either <strong>of</strong> the<br />

level <strong>of</strong> exposure or <strong>of</strong> the level <strong>of</strong> effect (Figure 2.5). For example, the<br />

level <strong>of</strong> urinary mercapturic acid excretion could assess the potential health<br />

hazard <strong>of</strong> an occupational exposure situation (Henderson et al., 1989).<br />

In practice, a complete view on the relation between toxicokinetics and<br />

toxicodynamics has not been elucidated for a single chemical up to now.<br />

Occupational monitoring methods all have their specific values based on<br />

their selectivity, sensitivity, validity and logistics and should therefore be<br />

used complementary to each other. All methods operate on the continuum<br />

from exposure to effect, the limits between which occupational toxicology<br />

studies operate.


Figure 2.5 Occupational monitoring methods and their relation to exposure versus<br />

effect assessment and to toxicokinetic and toxicodynamic processes. (Adapted from<br />

Henderson et al., 1989).<br />

Glutathione conjugation products as biomarkers<br />

In principle, GSH-conjugation derived metabolites can be used as a<br />

biomarker <strong>of</strong> internal dose. Glutathione (GSH), a tripeptide consisting <strong>of</strong><br />

the amino acids glycine, cysteine and -glutamine, plays an important role<br />

in the detoxification <strong>of</strong> potentially electrophilic chemicals or metabolites. In<br />

contrast, toxification via GSH-conjugation, for example <strong>of</strong> 1,2dibromoethane,<br />

hexachlorobutadiene, benzyl- and allylisothiocyanate has<br />

also been reported. β-lyase dependent bioactivation <strong>of</strong> cysteine-conjugates,<br />

derived from the initially formed GSH-conjugates, sometimes resulted in<br />

the formation <strong>of</strong> new reactive intermediates which are responsible for<br />

carcinogenic, mutagenic and other toxicological effects (Vermeulen, 1989;<br />

Van Welie et al., 1992).<br />

The initial step in GSH-conjugation is reaction <strong>of</strong> the nucleophilic<br />

sulphhydryl with electrophilic centers <strong>of</strong> a chemical. GSH-conjugation is<br />

catalysed by a family <strong>of</strong> glutathione S-transferase (GST) enzymes. A wide<br />

range <strong>of</strong> chemicals can be handled by this enzyme system due to the

existence <strong>of</strong> a large number <strong>of</strong> isoenzymes with different, though<br />

overlapping, substrate selectivity. The final detoxification capacity through<br />

GSH and GST enzymes <strong>of</strong> an organism depends on endogenous factors<br />

such as tissue distribution, genetic deficiencies, aging and hormonal<br />

influences and on exogenous factors such as sensitivity to inhibition and<br />

induction <strong>of</strong> GSTs (Vermeulen, 1989; Van Welie et al., 1992).<br />

GSH-conjugates normally are not excreted unchanged in urine or faeces.<br />

Catabolism <strong>of</strong> the GSH-conjugates results in the formation and excretion<br />

<strong>of</strong> a variety <strong>of</strong> sulphur containing metabolites, among which thioethers and<br />

mercapturic acids (S-substituted N-acetyl-cysteine conjugates) belong to the<br />

most important. The mercapturic acid pathway is shown in Figure 2.6.<br />

Thioethers in human studies<br />


Figure 2.6 Schematic representation <strong>of</strong> the mercapturic acid pathway: GSHconjugation<br />

with an electrophilic chemical (RX) and the biosynthesis to a<br />

mercapturic acid. E1: glutathione S-transferase, E2: -glutamyltranspeptidase, E3:<br />

cysteinylglycinase and aminopeptidase, E4: cysteine conjugate N-acetyltransferase,<br />

E5: N-deacetylase.<br />

Several years ago, Seutter-Berlage et al. proposed the appearance <strong>of</strong><br />

thioethers such as mercapturic acids (R-S-R′), mercaptans (R-SH) and<br />

disulfides (R-S-S-R′) in urine as an indicator <strong>of</strong> exposure to potentially<br />

alkylating chemicals. The thioether assay is an aselective assay to detect<br />

metabolic end-products excreted in urine <strong>of</strong> (non)occupational exposure to<br />

various electrophilic chemicals. It includes three steps, namely: (i)<br />

extraction, (ii) alkaline hydrolysis and (iii) derivatization, subsequently<br />

followed by spectrophotometric analysis at 412 nm. The thioether assay


Figure 2.7 Urinary excretion <strong>of</strong> thioethers (mmol SH/mol creatinine), <strong>of</strong> applicators<br />

exposed to 3.8 (− −), 9.8 (− −) and 18.9 (− −) mg m −3 8-h TWA (Z+E)-1,3dichloropropene<br />

in respiratory air, respectively. Darker shaded areas indicate<br />

exposure periods.<br />

was first applied to compare thioether excretions in urine <strong>of</strong> employees <strong>of</strong> a<br />

chemical plant. Highest thioether excretions were found in rubber workers<br />

and radial tyre builders when compared with clerks, plastic monomer<br />

mixers and footwear preparers. Recently, urinary thioether excretion was<br />

related to the occupational respiratory exposure <strong>of</strong> applicators in the Dutch<br />

flower-bulb culture to 1,3-dichloropropene (DCP) (Van Welie et al.,<br />

1991a). Instead <strong>of</strong> a discrete comparison <strong>of</strong> thioether excretion with<br />

exposed versus non-exposed groups, in this study thioether excretion was<br />

related to a continuous scale <strong>of</strong> airborne DCP concentrations.<br />

Significant linear relations between respiratory exposure to DCP and postminus<br />

preshift thioether concentration and cumulative thioether excretion<br />

were found. The urinary excretion <strong>of</strong> DCP-thioethers followed first-order<br />

elimination kinetics (Figure 2.7) with half-lives <strong>of</strong> elimination <strong>of</strong> 8.0±2.5 h<br />

(n=5) based on urinary excretion rates and 9.5±3.1 h (n=5) based on<br />

creatinine excretion. The elimination half-lives <strong>of</strong> the thioethers were<br />

almost two fold higher when compared to the half-lives <strong>of</strong> elimination <strong>of</strong><br />

the mercapturic acids <strong>of</strong> Z-and E-1,3 dichloropropene. This illustrates the<br />

main problem <strong>of</strong> urinary thioethers, viz. high background levels originating<br />

from endogenous or exogenous sources, such as smoking and diet (e.g.<br />

horse radish, onion and garlic).

Mercapturic acids in human studies<br />

Mercapturic acids, S-substituted N-acetyl-L-cysteine S-conjugates, in urine<br />

can be used as biomarkers <strong>of</strong> internal dose <strong>of</strong> electrophilic xenobiotics.<br />

Mercapturic acids are metabolic end products <strong>of</strong> GSH-conjugation <strong>of</strong><br />

various potentially electrophilic chemicals (Figure 2.6). The first<br />

mercapturic acids were identified in 1879 as sulphur containing<br />

metabolites after administration <strong>of</strong> bromobenzene to dogs (see references in<br />

Vermeulen, 1989). Since then mercapturic acids from many chemicals have<br />

been identified and these types <strong>of</strong> urinary metabolites have been used in<br />

biotransformation, biological monitoring and toxicological studies<br />

(Vermeulen, 1989; Van Welie et al., 1992).<br />

Commercial availability <strong>of</strong> reference compounds and the development <strong>of</strong><br />

a number <strong>of</strong> different analytical techniques attributed to the popularity <strong>of</strong><br />

mercapturic acids in biological monitoring studies during the last few<br />

years. Urinary excretion <strong>of</strong> the stereoisomeric mercapturic acids <strong>of</strong> Z- and<br />

E-1,3-dichloropropene, a soil fumigant frequently used in agriculture,<br />

proved to be a suitable biomarker for the exposure to both isomers in man.<br />

Strong correlations were observed between 8-h time weighted average<br />

exposure to Z- and E-DCP and complete cumulative excretion <strong>of</strong> N-acetyl-<br />

S-(Z- and E-3-chloropropenyl-2)-L-cysteine in urine. N-acetyl-S-<br />

(cyanoethyl)-L-cysteine was proposed as biomarker <strong>of</strong> exposure to<br />

acrylonitrile. The best correlation between uptake <strong>of</strong> acrylonitrile via the<br />

lungs and excretion <strong>of</strong> the cyanoethyl mercapturic acid in urine was<br />

obtained in samples collected between the sixth and the eighth hour after<br />

the beginning <strong>of</strong> exposure (Jakubowoski et al., 1987). The phenyl<br />

mercapturic acid <strong>of</strong> benzene was regarded as a useful biomarker <strong>of</strong><br />

exposure below 1 ppm <strong>of</strong> workers in a chemical production plant<br />

(Stommel et al., 1989). The use <strong>of</strong> certain foodstuffs and drugs may also<br />

give rise to the excretion <strong>of</strong> mercapturic acids. Consumption <strong>of</strong> cabbage<br />

and horse radish for example gave rise to increased thioether excretion.<br />

Consumption <strong>of</strong> garlic and onions resulted in the excretion <strong>of</strong> N-acetyl-S-<br />

(allyl- and 2-carboxypropyl)-L-cysteine in urine (Van Welie et al., 1992).<br />

The hypnotic drug ( α-bromo-isovalerylurea<br />

also gave rise to the excretion<br />

<strong>of</strong> two diastereomeric α-bromoisovalerylurea<br />

mercapturic acid conjugates<br />

in urine (Mulders et al., 1993). S-Phenyl mercapturic acid was present in<br />

urine <strong>of</strong> groups <strong>of</strong> smokers and non-smokers, not exposed to benzene, in<br />

concentrations <strong>of</strong> 4.0±4.0 µg g −1 creatinine (Stommel et al., 1989).<br />

Toxicokinetics<br />


Knowledge about the toxicokinetics <strong>of</strong> mercapturic acids is necessary to<br />

develop optimal sampling strategies in occupational studies. Urinary<br />

excretion rates <strong>of</strong> mercapturic acids theoretically may reflect the rates <strong>of</strong>


Figure 2.8 Urinary excretion (– –=Z, − −=E) and cumulative excretion (– –=Z,<br />

− −=E) <strong>of</strong> Z- and E-DCP-MA <strong>of</strong> an applicator due to an 8-h TWA respiratory<br />

exposure to 2.32 mg m −3 Z-DCP and 1.73 mg m −3 E-DCP. In (A) the mercapturic acid<br />

excretion rate is depicted and in (B) the mercapturic acid excretion based on<br />

creatinine excretion.<br />

elimination <strong>of</strong> the parent compounds from blood and can be used to<br />

calculate the (complete) cumulative excretion <strong>of</strong> mercapturic acids related<br />

to exposure. By knowing an individual’s mercapturic acid excretion rate,<br />

the contribution to urinary mercapturic acid excretion <strong>of</strong> the day under<br />

study on succeeding day(s) can be calculated. The contributions <strong>of</strong> previous<br />

days <strong>of</strong> exposure can also be used to correct the mercapturic acid excretion<br />

<strong>of</strong> the exposure day under study. The urinary half-life <strong>of</strong> elimination is<br />

inversely proportional to the elimination rate constant. The urinary halflife<br />

<strong>of</strong> both mercapturic acids <strong>of</strong> Z- and E-DCP in man was ca. 5 h<br />

(Figure 2.8) and they were not significantly different, i.e. 5.0±1.2 h for Z-<br />

DCP-MA and 4.7±1.3 h for E-DCP-MA. Strong corre-lations (r≥0.93) were<br />

observed between respiratory 8 h time weighted average (TWA) exposure<br />

to Z- and E-DCP and complete cumulative urinary excretion <strong>of</strong> Z- and E-<br />

DCP-MA. There is still a lack <strong>of</strong> knowledge about the magnitude <strong>of</strong> the


intra- and inter-individual differences in GSH-conjugation and mercapturic<br />

acid excretion. Factors causing these differences are sex, stress, diet, age,<br />

enzyme induction and inhibition, pathology and genetic variability. Apart<br />

from these factors the presence or absence <strong>of</strong> glutathione S-transferases<br />

(GSTs) or GST activity in different persons is <strong>of</strong> special interest in relation<br />

to urinary mercapturic acid excretion. The most intriguing factor known in<br />

this context is the human genetic polymorphism <strong>of</strong> mu-class GSTs. The<br />

GST isoenzyme µ is expressed only in approximately 60% <strong>of</strong> the human<br />

population. Mu-class GST isoenzymes showed a high specific activity<br />

towards for example styrene-7,8-oxide and benzo(a)pyrene-4,5dihydrodiol-4,5-oxide<br />

and E- and Z-DCP. Genetic polymorphism <strong>of</strong> muclass<br />

GSTs was postulated as a determinant in the excretion <strong>of</strong> the<br />

mercapturic acids <strong>of</strong> Z- and E-DCP in occupationally exposed applicators.<br />

However, between mu-class positive (n=9) and mu-class<br />

Table 2.2 Urinary excretion levels, urinary ratios and half-lives <strong>of</strong> elimination <strong>of</strong> Zand<br />

E-DCP mercapturic acids <strong>of</strong> mu-class positive and mu-class negative<br />

individuals a<br />

a Urinary excretion level represents the cumulative excretion <strong>of</strong> Z- and E-DCP-MA<br />

in 0–36 h urine, corrected for the time weighted average 8-h exposure to Z- and E-<br />

DCP. Values are expressed as means±SD for the number <strong>of</strong> individuals indicated in<br />

parentheses.<br />

b (mmol mercapturic acid)/(mmol DCP m −3 ).<br />

c Z-DCP-MA/E-DCP-MA<br />

d Half-life <strong>of</strong> elimination<br />

negative (n=3) applicators, neither a difference in urinary half-lives <strong>of</strong><br />

elimination nor in cumulative excretion <strong>of</strong> both mercapturic acids <strong>of</strong> Zand<br />

E-DCP was seen (Vos et al., 1991) (Table 2.2).<br />

α-Bromoisovalerylurea,<br />

a sedative and hypnotic drug, is a racemic drug<br />

which is also metabolized by GSH-conjugation. It was proposed as a<br />

model substrate to study the pharmacokinetics and stereoselectivity <strong>of</strong> GSHconjugation<br />

in humans. Stereoselective mercapturic acid formation <strong>of</strong> Rand<br />

S-α-bromoisovalerylurea was seen in in vitro studies with purified GST<br />

isoenzymes and in vivo in rat and man. In humans, a pronounced<br />

stereoselectivity in urinary mercapturic acid excretion was observed. Of an<br />

oral dose <strong>of</strong> R- and S-α-bromoisovalerylurea, 22.5±4.3 and 5.7±1.6% was<br />

excreted as mercapturic acid in 24 h, respectively. The half-lives <strong>of</strong><br />

elimination <strong>of</strong> both diastereoisomeric mercapturic acids were 1.5±0.4 and<br />

3.1±1.3 h, respectively. Both the pharmacokinetics <strong>of</strong> α-bromoisovaleryl


Figure 2.9 Proposed biotransformation pathway <strong>of</strong> etridiazol leading to 5-ethoxy-1,<br />

2,4-thiadiazole-3-carboxylic acid (ET-CA) and N-acetyl-S-(ethoxy-1,2,4thiadiazol-3-yl-methyl)-L-cysteine<br />

(ET-MA) in rat and humans. Unidentified<br />

intermediates are presented between brackets ([...]). GSH: glutathione, MAP:<br />

mercapturic acid pathway.<br />

ureas and their stereoselectivity, however, were not found to be different for<br />

subjects who were GSH S-transferase class mu deficient and subjects who<br />

were not (Mulders et al., 1993).<br />

Disposition <strong>of</strong> etridiazol<br />

Etridiazol (Aaterra; 5-ethoxy-3-trichloromethyl-l,2,4-thiadiazole<br />

(Figure 2.9)) is an agricultural fungicide used to control phycomycetous<br />

fungi in, for example, plants, tomatoes, cucumbers, cauliflowers and<br />

celery. Concerning external exposure <strong>of</strong> applicators (e.g. greenhouse<br />

handgunners and foggers) it has been concluded that exposure may occur<br />

through inhalation and dermal absorption. For the purpose <strong>of</strong> the<br />

development <strong>of</strong> a biomonitoring assay disposition studies were performed<br />

recently in rats and human volunteers (Van Welie et al., 1991c). Two<br />

metabolites, 5-ethoxy-l,2,4-thiadiazole-3-carboxylic acid (ET-CA) and a<br />

mercapturic acid, N-acetyl-S-(5-ethoxy-l,2,4-thiadiazol-3-yl-methyl)-Lcysteine<br />

(ET-MA) were identified as new metabolites. Based on a<br />

preliminary toxicokinetic study, the urinary excretion <strong>of</strong> the former<br />

metabolite amounted to 22±9% <strong>of</strong> an oral dose <strong>of</strong> etridiazol (while ET-MA<br />

and unchanged etridiazol were less than 1 % <strong>of</strong> the dose), ET-CA was<br />

proposed as a possible biomarker <strong>of</strong> exposure to this fungicide.

1,1,2-Trichloroethylene<br />


The solvent properties <strong>of</strong> 1,1,2-trichloroethylene (TRI) have resulted in its<br />

widespread use in metal degreasing and a wide variety <strong>of</strong> other industrial<br />

applications. TRI has now been in common use for more than 50 years.<br />

During this period <strong>of</strong> time, workers have been exposed to a wide range <strong>of</strong><br />

concentrations, in some cases for periods <strong>of</strong> 25 years or longer. This has<br />

allowed the compilation <strong>of</strong> a great data base about the effects <strong>of</strong> TRI on<br />

human health. Moreover, information has been supplemented by<br />

numerous studies in experimental animals.<br />

Epidemiological studies on more than 15000 individuals with a followup<br />

<strong>of</strong> more than 25 years have shown no evidence <strong>of</strong> an association<br />

between human exposure to TRI and increased incidence <strong>of</strong> cancer or<br />

cancer mortality. However, several <strong>of</strong> these studies had more or less serious<br />

shortcomings. A summary <strong>of</strong> effects related to TRI and/or TRI-related<br />

metabolism is given in Table 2.3. These and other data are taken from<br />

Goeptar et al., 1995a.<br />

An increased incidence <strong>of</strong> lung tumors has been reported in female<br />

B 6C 3F 1 and male Swiss mice exposed to TRI by inhalation. The effect was<br />

not observed in male B 6C 3F 1 nor in female Swiss mice nor in rats. This<br />

apparent strain-, sex- and lung-specific response fails to resolve the issue <strong>of</strong><br />

whether or not TRI is a carcinogenic hazard to man. Mechanistic studies<br />

on mouse lung tumor formation have explained the sex and species<br />

differences. In this context, chloral formation (Figure 2.10) in Clara cells,<br />

containing relatively high cytochrome P-450 concentrations, has been<br />

identified to be responsible for the development <strong>of</strong> mouse lung tumors.<br />

Importantly, lung tumors have not been found in humans after long-term<br />

occupational exposure in TRI.<br />

TRI causes an increase in the incidence <strong>of</strong> liver cancer in both sexes <strong>of</strong><br />

B 6C 3F 1 and Swiss mice following either gavage or inhalatory exposure, but<br />

not in NMRI and Ha: ICR mice nor in rats. A rodent specific link between<br />

peroxisome proliferation, DNA synthesis, inhibition <strong>of</strong> intercellular<br />

communication and cancer (Table 2.3) suggests that these responses are the<br />

basis <strong>of</strong> the hepatocarcinogenicity induced by TRI. The identification <strong>of</strong><br />

TCA in cancer bioassays as the responsible metabolite for these effects<br />

confirmed this hypothesis. However, when TCA was administered to both<br />

rats and mice, liver cancer was only observed in mice and not in rats. The<br />

reason for this species selectivity in liver effects is explained by the kinetic<br />

behavior <strong>of</strong> TRI and TCA in rodents. Both rats and mice have a considerable<br />

capacity to metabolize TRI to TCA and TCE, the maximal capacities being<br />

closely related to the relative surface areas rather than to their body<br />

weights. Oxidative metabolism <strong>of</strong> TRI in rats is linearly related to dose at<br />

lower dose levels, but it becomes saturated at higher dose levels. Thus, an<br />

important difference between rats and mice is the lower saturation


Table 2.3 Reported toxic effects related to TRI and/or TRI-derived metabolites<br />

a References: see review Goeptar et al., 1995b.<br />

n.d.: not determined.


Figure 2.10 Oxidative metabolism <strong>of</strong> TRI in the rodent and mammalian liver and<br />

the formation <strong>of</strong> metabolites which are excreted in the urine.<br />

concentration in the former species. The relevance <strong>of</strong> the mechanisms <strong>of</strong><br />

liver tumor formation in B 6C 3F 1 and Swiss mice for humans exposed to<br />

TRI has been assessed in studies comparing metabolic rates in mice, rats<br />

and humans. In contrast to the rat, the oxidative metabolism <strong>of</strong> TRI to<br />

TCA in humans is not limited by saturation. In this respect, humans<br />

resemble the mouse and might be able to produce sufficient TCA to induce<br />

peroxisome proliferation and consequently liver cancer. However, there are<br />

significant differences between mice and humans. First, humans metabolize<br />

approximately 60 times less TRI on a body weight basis than mice at<br />

similar exposure levels. Second, TCA has been shown to induce<br />

peroxisome proliferation in mouse hepatocytes but not in human<br />

hepatocytes (Table 2.3). Consequently, the combination <strong>of</strong> extensive<br />

oxidative metabolism <strong>of</strong> TRI to TCA and the ability <strong>of</strong> TCA to induce<br />

peroxisome proliferation appear to be unique to B 6C 3F 1 and Swiss mice.<br />

TRI-induced renal toxicity and tumors were found in Sprague-Dawley,<br />

Fischer 344 and Osborne-Mendel rats. These nephrocarcinogenic effects <strong>of</strong><br />

TRI were specific to male rats and were not seen in female rats nor in mice<br />

<strong>of</strong> either sex. 1,2-DCV-Cys, formed from TRI via the mercapturic acid<br />

pathway, has been identified as a likely metabolite involved in the observed<br />

renal toxicity and probably also in renal carcinogenicity in rats. TRI is<br />

metabolized by a minor pathway involving initial hepatic GSH-conjugation<br />

<strong>of</strong> TRI. The resulting DCV-G is further metabolized (Figure 2.11) and<br />

excreted in urine as two regioisomeric mercapturic acids, namely vicinal 1,<br />

2-DCV-Nac and geminal 2,2-DCV-Nac (Figure 2.11). 1,2-DCV-Cys (the<br />

precursors <strong>of</strong> 1,2-DCV-Nac) is a substrate for the renal L-cysteine S-


Figure 2.11 Possible routes <strong>of</strong> metabolism <strong>of</strong> S-(1,2-dichlorovinyl)glutathione (1,2-<br />

DCV-G). Steps are catalyzed by (a) -glutamyltransferase; (b) cysteinylglycine<br />

dipeptidase; (c) L-cysteine S-conjugate β-lyase; (d) L-cysteine S-conjugate Nacetyltransferase;<br />

(e) acylase.<br />

conjugate β-lyase and it is more mutagenic and cytotoxic than 2,2-DCV-<br />

Cys (the precursors <strong>of</strong> 2,2-DCV-Nac).<br />

The bioactivation <strong>of</strong> 1,2-DCV-Cys is without a doubt a crucial step in<br />

the onset <strong>of</strong> nephrotoxicity in the rat, although the precise biological<br />

mechanisms by which these metabolites exert their nephrocarcinogenic<br />

effects are not yet fully understood. A key aspect in the onset <strong>of</strong><br />

nephrocarcinogenicity in rats, however, is that it will not occur in the<br />

absence <strong>of</strong> nephrotoxicity. This suggests that the alkylating effects <strong>of</strong> the<br />

reactive metabolites (most likely thioketenes) derived from bioactivation <strong>of</strong><br />

1,2-DCV-Cys by β-lyase may not be sufficient to cause kidney tumors. The<br />

specific activity <strong>of</strong> β-lyase, the key enzyme involved in the bioactivation <strong>of</strong><br />

DCV-Cys isomers, is similar in humans to that in the mouse and only 10%<br />

<strong>of</strong> that in the rat. Moreover, human TRI metabolism via the mercapturic<br />

acid pathway resembles that <strong>of</strong> the mouse. It is, therefore, questionable<br />

whether humans are able to produce sufficient DCV-Cys isomers from TRI<br />

to cause first nephrotoxicity and then nephrocarcinogenicity. An important<br />

finding is also that the occurrence <strong>of</strong> nephrotoxicity and


nephrocarcinogenicity in the male rat is dose-dependent. More specifically,<br />

cytotoxic kidney damage is a feature <strong>of</strong> high continuous exposure to TRI<br />

over prolonged periods <strong>of</strong> time. This is unlikely to occur in humans during<br />

occupational exposure. In fact, TRI has been found not to be nephrotoxic<br />

in humans chronically exposed to low levels <strong>of</strong> TRI (50 mg m −3 ).<br />

Consequently, it is unlikely that the renal tumors which are seen in rats at<br />

nephrotoxic dose levels <strong>of</strong> TRI and which are related to β-lyase mediated<br />

bioactivation <strong>of</strong> 1,2-DCV-Cys, are relevant to human health hazards at<br />

reasonably foreseeable levels <strong>of</strong> exposure.<br />

Physiologically based toxicokinetic modeling <strong>of</strong> 1,3butadiene<br />

Physiologically based pharmaco(toxico)-kinetic models differ from the<br />

conventional compartmental models in that they are based to a large extent<br />

on the actual physiology <strong>of</strong> the organism. Instead <strong>of</strong> compartments defined<br />

largely by the experimental data themselves, actual organ and tissue groups<br />

are used with weights and blood flows from the literature (Bisch<strong>of</strong> and<br />

Brown, 1966). Instead <strong>of</strong> composite rate constants determined by fitting<br />

the actual experimental data, physical-chemical and biochemical constants<br />

<strong>of</strong> the compound are used. The result is a mode which predicts the<br />

qualitative behavior <strong>of</strong> the experimental time course without being based<br />

on it. Refinements <strong>of</strong> the model to incorporate additional insights gained<br />

from comparison with experimental data yields a model which can be used<br />

for quantitative extrapolations well beyond the range <strong>of</strong> experiments. In<br />

recent years several PBTK- and PBPK-models have been published: for<br />

methylene chloride, see Andersen et al., 1987; for a review see Leung et al.,<br />

1988; for 1,3-butadiene, see Evelo et al., 1993.<br />

The development <strong>of</strong> a PBTK/PBPK model can be divided into a number<br />

<strong>of</strong> steps: (a) inventory <strong>of</strong> physiological and toxicological behaviour <strong>of</strong> the<br />

compound, (b) mathematical description <strong>of</strong> the biochemical/(patho)<br />

physiological processes involved, (c) parameterization <strong>of</strong> the mathematical<br />

descriptions, (d) the construction <strong>of</strong> the model, (e) refinement and<br />

validation <strong>of</strong> the model and (f) use <strong>of</strong> the predictions and risk assessment.<br />

As an illustrative example <strong>of</strong> this approach the recently described PBTKmodeling<br />

<strong>of</strong> 1,3-butadiene disposition and toxicity might be used (Evelo et<br />

al., 1993). 1,3-Butadiene used for the production <strong>of</strong> styrene-butadiene<br />

rubber, is known amongst others to cause lung carcinogenicity. In the rat<br />

the carcinogenicity <strong>of</strong> 1,3-butadiene is less pronounced while the evidence<br />

for human carcinogenicity is inconclusive, Monoand di-epoxy-butadiene<br />

are reactive metabolites held responsible for this effect. Butadiene<br />

monoxide is formed by microsomal fractions <strong>of</strong> the lung and liver <strong>of</strong> several


Figure 2.12 Physiologically based toxicokinetic model for description <strong>of</strong> butadiene<br />

distribution and metabolism in mice, rats and humans. Gas exchange occurs in the<br />

alveoli <strong>of</strong> the lung. Metabolism occurs in both the alveolar and bronchial areas <strong>of</strong><br />

the lung and in the liver. Metabolic activity in the three other compartments is<br />

ignored (Evelo et al., 1993).<br />

species. There are, however, large interspecies differences in the lung vs<br />

liver activities: mice>rats>humans/monkeys.<br />

The PBTK model used to describe butadiene distribution and metabolism<br />

in mice, rats and humans is shown in Figure 2.12. Gas exchange is<br />

supposed to occur in the alveoli <strong>of</strong> the lung and metabolism in both the<br />

alveolar and bronchial areas <strong>of</strong> the lung and in the liver. By using the<br />

experimentally determined or estimated species-selective parameters for

volumes, masses and blood flows <strong>of</strong> different organs, partition coefficients<br />

<strong>of</strong> 1,3-butadiene between blood and organs/tissues and for metabolic<br />

capacities in liver and lung (bronchial and alveolar areas), accurate dosedependent<br />

simulations were performed for the uptake <strong>of</strong> 1,3-butadiene in<br />

mice and rats in gas-closed chambers. Moreover, with the resulting model<br />

the relative importance <strong>of</strong> lung metabolism as compared to metabolism in<br />

the liver was predicted for the three different species. Lung metabolism<br />

appeared to be much more important than liver metabolism in mice, this in<br />

contrast to the situation in the rat and humans. Moreover, at low exposure<br />

concentrations the relative importance <strong>of</strong> lung metabolism was predicted to<br />

increase in mice as a result <strong>of</strong> diminished saturation <strong>of</strong> metabolism in this<br />

species. It was concluded that the observed species differences in lung vs<br />

liver metabolism <strong>of</strong> 1,3-butadiene (mice>rat>human) and the tendency<br />

towards increased lung metabolism at low doses might rationalize the<br />

observed species differences in the lung carcinogenicity <strong>of</strong> 1,3-butadiene<br />

and this knowledge should be useful in the in vivo extrapolation from high<br />

dose to low dose risk assessments within one species as well as in<br />

interspecies risk assessment extrapolations.<br />

Conclusions<br />

In conclusion, a pr<strong>of</strong>ound knowledge <strong>of</strong> the biodisposition and the<br />

toxicokinetics <strong>of</strong> a toxic or potentially toxic chemical is <strong>of</strong> utmost<br />

importance to the design and interpretation <strong>of</strong> laboratory assessments <strong>of</strong><br />

toxicity, to explain interspecies differences in toxicities and to extrapolate<br />

more reliably from animal experiments to man in the process <strong>of</strong> risk<br />

assessment. This also holds true for the design for proper biological<br />

monitoring procedures and for the interpretation <strong>of</strong> the results in terms <strong>of</strong><br />

potential health risks <strong>of</strong> exposure to chemicals. Apart from traditional<br />

compartment-based toxicokinetic approaches, more recent physiologicallybased<br />

toxicokinetics modeling approaches have distinct advantages for the<br />

above-mentioned purposes.<br />

References<br />


ACGIH, 1990, in 1990–1991 Threshold limit values for chemical substances and<br />

physical agents and biological exposure indices, American Conference <strong>of</strong><br />

Governmental <strong>Industrial</strong> Hygienists, No. 0205.<br />


R.H., 1987, Physiologically-based pharmacokinetics and the risk assessment for<br />

methylene chloride, Toxicol. Appl. Pharmacol., 87, 185–205.<br />

ARIENS, E.J. and SIMONIS, M.A., 1980, in BREIMER, D.D. (Ed.) Towards better<br />

Safety <strong>of</strong> Drugs and Pharmaceutical Products, Amsterdam: Elsevier<br />

Biomedical Press.


BISCHOF, K.B. and BROWN, R.G., 1966, Drug distribution in mammals, Chem.<br />

Eng. Prog. Symp. Ser., 62(66), 33–45.<br />

CCRX 1989, in Metingen van radioactiviteit en xenobiotische st<strong>of</strong>fen in het<br />

biologische milieu in Nederland 1989 (in Dutch with English summary),<br />

Coördinatie-commissie voor de metingen van radioactiviteit en xenobiotische<br />

st<strong>of</strong>fen, Bilthoven: RIVM.<br />

COMMANDEUR, J.N.M. and VERMEULEN, N.P.E. 1991, Molecular and<br />

biochemical mechanism <strong>of</strong> chemically induced nephrotoxicity: a review, Chem.<br />

Res. Toxicol., 3, 171–94.<br />


1993, Physiologically based toxicokinetic modeling <strong>of</strong> 1,3-butadiene lung<br />

metabolism in mice becomes more important at low doses. Environ. Hlth<br />

Perspect., 101(6), 496–502 (no. 24).<br />

FARMER, P.B., NEUMANN, H.-G. and HENSCHLER, D., 1987, Estimation <strong>of</strong><br />

exposure <strong>of</strong> man to substances reacting covalently with macromolecules, Arch.<br />

Toxicol, 60, 251–60.<br />


P.J. VAN and VERMEULEN, N.P.E. 1995a, The metabolism and kinetics <strong>of</strong><br />

trichloroethylene in relation to toxicity and carcinogenicity. Relevance <strong>of</strong> the<br />

Mercapturic Acid Pathway, Chem. Res. Toxicol, 8, 3–21.<br />

GOEPTAR, A.R., SCHEERENS, H. and VERMEULEN, N.P.E., 1995b, Oxygen<br />

and xenobiotic reductase activities <strong>of</strong> cytochrome P450, Crit. Rev. Toxicol.,<br />

25, 25–65.<br />

GUENGERICH, F.P., 1994, Catalytic selectivity <strong>of</strong> human cytochrome P450<br />

enzymes: relevance to drug metabolism and toxicity, Toxicol. Lett., 70, 133–8.<br />

HENDERSON, R.F., BECHTOLD, W.E., BOND, J.A. and SUN, J.D., 1989, The use<br />

<strong>of</strong> biological markers in toxicology, Crit. Rev. Toxicol, 20, 65–82.<br />

HOOK, J.B. and HEWITT, W.R., 1986, Toxic responses <strong>of</strong> the kidney, in Klaassen,<br />

C.D., Doull, J. and Amdur, M.O. (Eds) Casarett and Doull’s <strong>Toxicology</strong>, pp.<br />

310–29, New York: Macmillan.<br />

JAKUBOWOSKI, M., LINHART, I., PIELAS, G. and KOPECKY, J., 1987, 2-<br />

Cyanoethylmercapturic acid (CEMA) in the urine as a possible indicator <strong>of</strong><br />

exposure to acrylonitrile, Brit. J. Ind. Med., 44, 834–40.<br />


VERMEULEN, N.P.E., 1993, Cytochromes P450: their active-site structure<br />

and mechanism <strong>of</strong> oxidation, Drug Metab. Rev., 25, 325–87.<br />

LEUNG, H.W., Ku, R.H., PAUSTENBACH, D.J. and ANDERSEN, M.E., 1988, A<br />

physiologically-based pharmacokinetic model for 2,3,7,8-tetrachlorodibenzo-pdioxin<br />

in C57BL/6J and DBA/2J mice, Toxicol. Lett., 42, 15–28.<br />


BREIMER, D.D. and MULDER, G.J., 1993, Characterization <strong>of</strong> glutathione<br />

conjugation in humans: stereoselectivity in plasma elimination<br />

pharmacokinetics and urinary excretion <strong>of</strong> (R)- and (S)-2-bromoisovalerylurea<br />

in healthy volunteers. Clin. Phar. Ther., 53, 49–58.<br />

POTT, P. 1795, Chirurgical observations relative to the cataract, the polypus <strong>of</strong> the<br />

nose, the cancer <strong>of</strong> the scrotum, the different kinds <strong>of</strong> ruptures and the<br />

mortification <strong>of</strong> the toes and feet, in Haes, Clarke and Collins (Eds) National<br />

Cancer Institute Monograph, 1962, Vol 10, pp. 7–13, London.



P.T.H., 1979, Urinary mercapturic acid excretion as a biological parameter <strong>of</strong><br />

exposure to alkylating agents, Int. Arch. Occup. Environ. Hlth, 39, 45–51.<br />

SIPES, I.G. and GANDOLFI, A.J., 1986, Biotransformation <strong>of</strong> chemicals, in<br />

Klaassen, C.D., Doull, J. and Amour, M.O. (Eds) Casarett and Doull’s<br />

<strong>Toxicology</strong>, pp. 64–98, New York: Macmillan.<br />


and NORPOTH, K., 1989, Determination <strong>of</strong> S-phenylmercapturic acid in the<br />

urine an improvement in the biological monitoring <strong>of</strong> benzene exposure,<br />

Carcinogenesis, 10, 279–82.<br />


VERMEULEN, N.P.E., 1991a, Thioether excretion in urine <strong>of</strong> applicators<br />

exposed to 1,3-dichloropropene: a comparison with urinary mercapturic acid<br />

excretion, Brit. J. Ind. Med., 48, 492–8.<br />


and VERMEULEN, N.P.E., 1991b, Inhalation exposure to 1.3dichloropropene<br />

in the Dutch flower-bulb culture. Part II. Biological<br />

monitoring by measurement <strong>of</strong> urinary excretion <strong>of</strong> two mercapturic acid<br />

metabolites, Arch. Environ. Contam. Toxicol, 20, 6–12.<br />


E. 1991c, Identification and quantitative determination <strong>of</strong> a carboxylic and a<br />

mercapturic acid metabolite <strong>of</strong> etridiazole in urine <strong>of</strong> rat and man. Potential<br />

tools for biological monitoring. Arch. Toxicol., 65, 625–32.<br />


SITTERT, N.J., 1992, Mercapturic acids, protein adducts, and DNA adducts<br />

as biomarkers <strong>of</strong> electrophilic chemicals, Crit. Rev. Toxicol., 22, 271–306.<br />

VERMEULEN, N.P.E., 1989, Analysis <strong>of</strong> mercapturic acids as a tool in<br />

biotransformation, biomonitoring and toxicological studies. TiPS, 10, 177–81.<br />



VAN WELIE, R.T.H., 1990, Molecular mechanisms in toxicology and drug<br />

design, in Claassen, V. (Ed.) Vol. 13, pp. 253–71, Trends in Drug Research,<br />

Amsterdam: Elsevier.<br />

Vos, R.M.E., VAN WELIE, R.T.H., PETERS, W.H.M., EVELO,<br />


P.J., 1991, Genetic deficiency <strong>of</strong> human class mu glutathione S-transferase<br />

isoenzymes in relation to the urinary excretion <strong>of</strong> the mercapturic adds <strong>of</strong> Zand<br />

E-1,3-dichloropropene. Arch. Toxicol., 65, 95–9.<br />

WARHOLM, M., JENSSON, H., TAHIR, M.K. and MANNERVIK, B., 1986,<br />

Purification and characterization <strong>of</strong> three distinct glutathione S-transferases<br />

from mouse liver, Biochemistry., 25, 4119–25.<br />

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activities and their relevance for the practice <strong>of</strong> occupational health, Int. Arch.<br />

Occup. Environ. Hlth, 57, 249–57.

3<br />

Metabolic Activation <strong>of</strong> <strong>Industrial</strong> Chemicals and<br />

Implications for Toxicity<br />


Leiden University, Leiden<br />

Introduction<br />

In the toxicity <strong>of</strong> industrial chemicals bioactivation (Anders, 1985) plays an<br />

important role. Obviously, its importance depends on the structure <strong>of</strong> the<br />

chemical as well as the toxic effect considered. Thus, inorganic compounds<br />

in general will not require bioactivation: metal salts or oxides will usually<br />

cause toxicity in the form in which they are taken up. However, even these<br />

chemicals may require further metabolism for maximum toxicity in the<br />

body: inorganic mercury may be converted to an organic form<br />

(methylmercury), and nitrate may be reduced to nitrite. It is also possible<br />

that in vivo complexes are being formed, such as between heavy metals<br />

ions and the protein, metallothionein, which may be more toxic (or cause<br />

more organ-selective toxicity) than the original, uncomplexed compound<br />

(Wang et al., 1993).<br />

Bioactivation thus mostly concerns the conversion <strong>of</strong> organic chemicals<br />

to more toxic products. On one hand this may result in stable metabolites<br />

that better fit a receptor binding site, resulting in (in principle) reversible<br />

interactions (Mulder, 1992). On the other hand, the metabolites may be<br />

quite reactive, resulting in essentially irreversible effects which are <strong>of</strong><br />

particular concern when they can escape correction, such as neoplasms or<br />

sensitization.<br />

Mechanisms <strong>of</strong> bioactivation<br />

<strong>Industrial</strong> chemicals have widely different structures. Often the preparations<br />

used contain a variable degree <strong>of</strong> impurities, or are mixtures. In this<br />

chapter only the toxicity <strong>of</strong> pure chemicals will be discussed; obviously<br />

when several compounds are present at the same time in a reaction mix or<br />

a commercial product, the final toxicity may be the result <strong>of</strong> complex<br />

interactions between the substituents, which may cause the toxicity to be<br />

more severe (but also much less serious) than expected.

The bioactivation to reactive intermediates by oxidative, cytochrome<br />

P450-mediated metabolism has been extensively studied. So much so, that<br />

it is <strong>of</strong>ten overlooked that conjugation reactions may similarly convert<br />

stable compounds into reactive, electrophilic metabolites (Anders and<br />

Dekant, 1994). This is <strong>of</strong> some practical importance, because many rapid<br />

in vitro toxicity screening tests, e.g. for genotoxicity, include only oxidative<br />

biotransformation capacity (microsomal fractions plus NADPH). In such<br />

screening systems the possibility that, for example, glucuronidation,<br />

sulfation or glutathione conjugation may activate a chemical is not<br />

assessed. Examples <strong>of</strong> bioactivation <strong>of</strong> industrial chemicals by glutathione<br />

conjugation are various halogenated hydrocarbons, while in 2naphthylamine<br />

toxicity glucuronidation may play a role. All in all,<br />

however, little information is available on the role <strong>of</strong> conjugation. As a<br />

consequence, it is unclear at present whether conjugation reactions are <strong>of</strong><br />

major concern for bioactivation <strong>of</strong> industrial chemicals in general. It<br />

certainly seems worth while for reasons more than just scientific curiosity<br />

to include conjugation reactions in test systems. This can be done by using,<br />

for example, intact hepatocytes (or other cells), or by using a mix <strong>of</strong><br />

cosubstrates for conjugation in combination with an S9 fraction (consisting<br />

<strong>of</strong> both cytosol and microsomal fraction). UDP glucuronic acid, a sulfate<br />

activating system, glutathione, acetyl-CoA and S-adenosylmethionine<br />

would cover the major conjugation reactions.<br />

A role <strong>of</strong> bioactivation in the toxicity <strong>of</strong> many chemicals has been<br />

demonstrated. Chemical groups that <strong>of</strong>ten are involved in mutagenic or<br />

carcinogenic effects have been identified (‘alerting groups’). However, as yet<br />

it is still impossible to predict with certainty the carcinogencity <strong>of</strong> a<br />

compound based only on its chemical structure, although a panel <strong>of</strong><br />

experts can make quite good guesses (Wachsman et al., 1993).<br />

In this chapter some <strong>of</strong> the major issues will be illustrated by the<br />

examples vinyl chloride, styrene (versus styrene oxide), benzene,<br />

dichloromethane, chlor<strong>of</strong>orm, 1,2-dibromoethane and 2-naphthylamine.<br />

Vinyl chloride<br />

G.J.MULDER 37<br />

High exposure <strong>of</strong> workers to vinyl chloride in the past has led to the<br />

realization that it may cause neoplasms in man, in particular<br />

haemangiosarcomas in the liver. Vinyl chloride is a genotoxic compound<br />

that acts as initiator <strong>of</strong> various types <strong>of</strong> tumors (Swaen et al., 1987).<br />

The major routes <strong>of</strong> bioactivation <strong>of</strong> vinyl chloride are shown in<br />

Figure 3.1. The most important first step is oxidation by (a) cytochrome<br />

P450 species, resulting in a rather reactive epoxide, which readily<br />

rearranges to chloroacetaldehyde. This may bind to DNA bases, especially<br />

the N6 <strong>of</strong> adenosine or the N4 <strong>of</strong> cytidine, yielding N-ethenoadducts.<br />

Glutathione provides protectionbecause it traps the reactive intermediates


Figure 3.1 Bioactivation <strong>of</strong> vinylchloride.<br />

formed from vinyl chloride. Furthermetabolism <strong>of</strong> such conjugates leads to<br />

urinary products that can be used tomonitor vinyl chloride exposure in<br />

workers (Guengerich, 1992).<br />

The compound is mutagenic in many in vitro test systems, which require<br />

bioactivation by a microsomal preparation with co-factors for cytochrome<br />

P450. Whether other toxic effects that have been associated with vinyl<br />

chloride exposure in man, such as Raynauds syndrome or acro-osteolysis,<br />

also require bioactivation <strong>of</strong> vinyl chloride is unknown. In addition to its<br />

DNA adduct forming capacity, vinyl chloride also binds covalently to thiol<br />

groups in proteins. It is conceivable that such binding in specific cell types<br />

might lead to non-carcinogenic defects in organ functions.<br />

Styrene and styrene oxide<br />

Styrene metabolism and bioactivation are very similar to that <strong>of</strong> vinyl<br />

chloride: epoxidation by cytochrome P450 is the pathway <strong>of</strong> toxification<br />

(Figure 3.2). It can be detoxified by epoxide hydrolase and glutathione<br />

transferase activity. Mandelic acid excretion in urine can be used for<br />

exposure monitoring in man. Styrene oxide is a direct mutagen in several in<br />

vitro mutagenesis systems and it readily reacts with DNA in vitro.<br />

However, when animals are exposed to styrene in vivo very little if any<br />

DNA binding is observed. Moreover, styrene is not carcinogenic in animal<br />

experiments, although it is a (weak) mutagen in vitro, after bioactivation<br />

(Bond, 1989; Ecetoc, 1992). The explanation most likely is that the styrene

Figure 3.2 Bioactivation <strong>of</strong> styrene.<br />

Figure 3.3 Bioactivation <strong>of</strong> chlor<strong>of</strong>orm.<br />

oxide, generated in vivo inside a cell is such a good substrate for the phase<br />

2 enzymes, epoxide hydrolase and glutathione transferase, that virtually<br />

immediately upon its synthesis, it is further metabolized. Thus, presumably<br />

the build-up <strong>of</strong> an effective concentration in vivo is prevented. Whether<br />

other toxicity <strong>of</strong> styrene in, for example, oesophagus, stomach or<br />

forestomach is related to covalent binding <strong>of</strong> styrene oxide to protein thiol<br />

groups in those tissues is unclear at present.<br />

Styrene is an example <strong>of</strong> a compound <strong>of</strong> which the metabolism<br />

completely goes through a reactive intermediate (the epoxide); yet it does<br />

not cause the cancer that might be expected from its highly mutagenic<br />

metabolite. Accumulation <strong>of</strong> enough <strong>of</strong> this epoxide inside the cells for a<br />

detectable genotoxic effect may require a dose which is acutely toxic, and<br />

therefore can never be tested.<br />

Chlor<strong>of</strong>orm<br />

G.J.MULDER 39<br />

Chlor<strong>of</strong>orm is acutely toxic in the liver and the kidney. This is the result <strong>of</strong><br />

formation <strong>of</strong> a reactive intermediate (Figure 3.3), phosgene, which binds


avidly to thiol and amine groups in protein. In mice the kidney toxicity is<br />

much more pronounced in males than in females; this sex-difference is due<br />

to the much higher activity <strong>of</strong> the bioactivating cytochrome P450 species in<br />

male mouse kidney than in the females (Pohl et al., 1984). Chlor<strong>of</strong>orm also<br />

increased the tumor incidence in the liver and kidney in some experiments<br />

(Reitz et al., 1990), at dose levels which damaged these organs. However,<br />

there are no indications <strong>of</strong> mutagenicity or genotoxicity in in vitro or<br />

animal in vivo systems. Therefore, most likely the increased tumor<br />

frequency in animals is due to tissue toxicity, leading to increased cell<br />

turnover and a mitogenic stimulus. This is an important distinction, at least<br />

in some countries such as The Netherlands, because for such chemicals a<br />

threshold approach is allowed, whereas for initiating chemicals a linear<br />

extrapolation for carcinogenic risk is used.<br />

Benzene<br />

Benzene presents something <strong>of</strong> a mystery in the evaluation <strong>of</strong> its toxicity<br />

mechanism (Swaen et al., 1989). Exposure to high levels <strong>of</strong> benzene has<br />

been associated with leukaemia in man. However, in vitro it shows little<br />

genotoxicity, and it hardly generates DNA adducts when it is given even at<br />

high dose to animals. A candidate for DNA damage could have been the 1,<br />

4-dihy-droxybenzene (hydroquinone) metabolite, which, however, does not<br />

form DNA adducts readily. Recently a ring-opened metabolite, the<br />

trans,trans-muconic dialdehyde has been proposed as a possible reactive<br />

metabolite <strong>of</strong> benzene (Figure 3.4). Whether it really plays a role in<br />

benzene toxicity is unclear as yet (Kline et al., 1993).<br />

Dichloromethane<br />

Dichloromethane can be metabolized by two pathways, an oxidative and a<br />

conjugative route. Oxidation catalyzed by P450 yields carbon monoxide<br />

(Figure 3.5). The glutathione pathway generates a reactive intermediate,<br />

which is mutagenic and has been implicated in the hepatocarcinogenic<br />

effect <strong>of</strong> dichloromethane in mice. It could be shown that the human liver<br />

has a negligible activity <strong>of</strong> the glutathione transferase involved, so that the<br />

risk for hepatocarcinogenesis in man is virtually non-existent (Green et al.,<br />

1988; Reitz et al., 1989; Dankovic and Bailer, 1994). This example<br />

illustrates how insight into the mechanism <strong>of</strong> bioactivation enables a more<br />

reliable species extrapolation in terms <strong>of</strong> hazard and risk.<br />

1,2-Dibromoethane<br />

This compound can be conjugated with glutathione to form a reactive<br />

thiiranium ion which forms adducts with DNA. This is the reason for the

Figure 3.4 Possible route <strong>of</strong> bioactivation <strong>of</strong> benzene.<br />

Figure 3.5 Bioactivation <strong>of</strong> dichloromethane.<br />

carcinogenic and mutagenic effects <strong>of</strong> 1,2-dibromoethane (Inskeep et al.,<br />

1986).<br />

2-Naphthylamine<br />

G.J.MULDER 41<br />

2-Naphthylamine causes bladder tumors in the dog and man, but not in<br />

mice and rats. The most likely cause is a complicated interplay between<br />

glucuroni dation and urinary pH. In all four species 2-naphthylamine is Nhydroxylated<br />

and subsequently N-glucuronidated. The resulting metabolite


is excreted in urine. In man and dog the urine is slightly acidic, while in rat<br />

and mouse it is slightly alkaline. Under acidic conditions the glucuronide is<br />

hydrolyzed to generate the hydroxylamine in the bladder. In this case<br />

glucuronidation is not a bioactivation, but rather a targeting<br />

biotransformation: in man and dog the carcinogenic metabolite is targeted<br />

to the bladder, due to the (necessary!) acidic local pH (Kadlubar et al.,<br />

1981).<br />

Conclusions<br />

The above illustrates the importance <strong>of</strong> bioactivation in toxicity <strong>of</strong><br />

industrial chemicals. Is it possible to predict bioactivation from the<br />

structure? As outlined above, in some cases the compound contains<br />

structural elements which make bioactivation to a reactive intermediate<br />

quite likely. Whether it does play a role in toxicity then is still uncertain.<br />

Test systems to detect reactive intermediates depend on, for example, the<br />

availability <strong>of</strong> the radiolabeled compound; in fact, a very high specific<br />

radioactivity is required to detect low levels <strong>of</strong> binding. Alternatively,<br />

radiolabelled glutathione can be used for those reactive intermediates that<br />

readily bind to the thiol group <strong>of</strong> glutathione (Mulder and Le, 1988).<br />

Whether such systems can pick up every relevant toxic reactive<br />

intermediate remains to be seen.<br />

For extrapolation <strong>of</strong> one species to the other it is important to have<br />

insight into the metabolite that is responsible for the toxicity. Therefore, it<br />

is more than just <strong>of</strong> academic interest to know the mechanism <strong>of</strong> toxicity in<br />

safety assessment <strong>of</strong> industrial chemicals. Unfortunately, it is <strong>of</strong>ten not easy<br />

to establish such a mechanism beyond reasonable doubt: it may require too<br />

many rats to feel comfortable about it if we would have to do this for every<br />

chemical used industrially!<br />

References<br />

ANDERS, M.W. (Ed.), 1985, Bioactivation <strong>of</strong> Foreign <strong>Compounds</strong>, Orlando, FL:<br />

Academic Press.<br />

ANDERS, M.W. and DEKANT, W., 1994, Conjugation-dependent Carcinogenicity<br />

and Toxicity <strong>of</strong> Foreign <strong>Compounds</strong>, Orlando, FL: Academic Press.<br />

BOND, J.A., 1989, Review <strong>of</strong> the toxicology <strong>of</strong> styrene, CRC Crit. Rev. Toxicol 19,<br />

227–49.<br />

DANKOVIC, D.A. and BAILER, A.J., 1994, The impact <strong>of</strong> exercise and<br />

intersubject variability on dose estimates for dichloromethane derived from a<br />

physiologically based pharmacokinetic model, Fund. Appl. Toxicol, 22, 20–5.<br />

ECETOC, 1992, Technical report No. 52, Styrene toxicology. Investigations on the<br />

potential for carcinogenicity, Brussels: Ecetoc.

G.J.MULDER 43<br />

GREEN, T., PROVAN, W.M., COLLINGE, D.C. and GUEST, A.E., 1988,<br />

Macro molecular interactions <strong>of</strong> inhaled methylene chloride in rats and mice,<br />

Toxicol. Appl. Pharmacol, 93, 1–10.<br />

GUENGERICH, F.R., 1992, Roles <strong>of</strong> the vinylchloride oxidation products 1chlorooxirane<br />

and 2-chloroacetaldehyde in the in vitro formation <strong>of</strong> etheno<br />

adducts <strong>of</strong> nucleic acid bases, Chem. Res. Toxicol, 5, 2–5.<br />

INSKEEP, P.B., KOGA, N.K., CMARIK, J.L. and GUENGERICH, F.P., 1986,<br />

Covalent binding <strong>of</strong> 1,2-dihaloalkanes to DNA, Cancer Res., 46, 2839–44.<br />


R.K. and MULDER, G.J., 1981, Alteration <strong>of</strong> urinary levels <strong>of</strong> the carcinogen,<br />

N-hydroxy-2-naphthylamine, and its N-glucuronide in the rat by control <strong>of</strong><br />

urinary pH, inhibition <strong>of</strong> metabolic sulfation, and changes in biliary excretion,<br />

Chem.-Biol. Interact. 33, 129–47.<br />


G., 1993, Identification <strong>of</strong> 6-hydroxy-trans,trans-2,4-hexadienoic acid, a novel<br />

ring-opened urinary metabolite <strong>of</strong> benzene, Environm. Hlth Perspect., 101,<br />

310–12.<br />

MULDER, G.J., 1992, Pharmacological effects <strong>of</strong> drug conjugates: is morphine 6glucuronide<br />

an exception? Trends Pharmacol. Sci., 13, 302–4.<br />

MULDER, G.J. and LE, C.T., 1988, A rapid simple in vitro screening test to detect<br />

reactive intermediates <strong>of</strong> xenobiotics. Toxicol. In Vitro, 2, 225–30.<br />

POHL, L.R., GEORGE, J.W. and SATOH, H., 1984, Strain and sex differences in<br />

chlor<strong>of</strong>orm-induced nephrotoxicity. Drug Metab. Disposit., 12, 304–7.<br />

REITZ, R.H., MENDRALA, A.L. and GUENGERICH, F.P., 1989, In vitro<br />

metabo-lism <strong>of</strong> methylene chloride in human and animal tissues, Toxicol.<br />

Appl. Pharmacol, 97, 230–46.<br />

REITZ, R.H., MENDRALA, A.L. and CONOLLY, R.B., 1990, Estimating the risk<br />

<strong>of</strong> liver cancer associated with human exposures to chlor<strong>of</strong>orm using PbPK<br />

modeling, Toxicol. Appl. Pharmacol., 105, 443–59.<br />

SWAEN, G.M.H. et al., 1987, A scientific basis for the risk assessment <strong>of</strong> vinyl<br />

chloride, Regul. Toxicol. Pharmacol, 7, 120–7.<br />

SWAEN, G.M.H. et al., 1989, Carcinogenic risk assessment <strong>of</strong> benzene in outdoor<br />

air, Regul. Toxicol. Pharmacol., 9, 175–85.<br />


TENNANT, R.W., 1993, Predicting chemical carcinogenesis in rodents,<br />

Environm. Hlth Perspect., 101, 444–5.<br />

WANG, X.P., CHAN, H.M., GOYER, R.A. and CHERIAN, M.G., 1993,<br />

Nephrotoxicity <strong>of</strong> repeated injections <strong>of</strong> cadmium-metallothionein in rats,<br />

Toxicol. Appl. Pharmacol., 119, 11–16.

4<br />

Sizing up the Problem <strong>of</strong> Exposure Extrapolation:<br />

New Directions in Allometric Scaling<br />


Director International Scientific Affairs, Servier Research and<br />

Development, Slough<br />

Introduction<br />

The evaluation <strong>of</strong> the safety <strong>of</strong> industrial chemicals requires the<br />

administration <strong>of</strong> a range <strong>of</strong> doses to test animals over periods <strong>of</strong> time and<br />

the extrapolation in some meaningful way to man. Various risk assessment<br />

models have been suggested which attempt to measure an uncertainty or<br />

safety factor which can be used to extrapolate to man to obtain an<br />

acceptable daily intake (ADI) (Dourson and Stara, 1983). Other<br />

approaches are also used, such as benchmark dose, the smallest dose which<br />

produces a statistical increase in toxicity over the background level (Crump,<br />

1984), or more frequently the LOEL, the lowest observed dose which<br />

produces an adverse effect, and NOEL, the highest dose at which no<br />

adverse effect is observed. There are difficulties in the interpretation <strong>of</strong><br />

these exposure margins since there is <strong>of</strong>ten little information on: (1) the<br />

slope or intensity <strong>of</strong> the effect, (2) species differences in the sensitivity, (3)<br />

the possibility <strong>of</strong> cumulative or irreversible toxicities, etc. But perhaps the<br />

most important weakness in these estimates is the lack <strong>of</strong> knowledge <strong>of</strong> the<br />

actual circulating levels <strong>of</strong> the chemical(s) in the different species. This<br />

problem is particularly pertinent for industrial chemicals and environmental<br />

pollutants where it may be unethical to administer doses <strong>of</strong> these<br />

compounds to volunteers which are sufficiently high to measure the<br />

kinetics. It is <strong>of</strong> special concern since it is well known that there are large<br />

interspecies differences in the clearance <strong>of</strong> chemicals and that comparison <strong>of</strong><br />

doses in animals, expressed simply in terms <strong>of</strong> mg kg −1 , provides little<br />

information as to the actual exposure likely to occur. This is not surprising<br />

since small animals have relatively faster blood flow and larger organs than<br />

man when expressed as a percentage <strong>of</strong> body weight, and consequently<br />

clearance is more rapid and circulating levels <strong>of</strong> the administered<br />

compound are lower than could be expected during toxicity testing<br />

(Campbell and Ings, 1988).<br />

However since most mammals share similar physiological and<br />

biochemical actions these differences in physiological rates and sizes for

most processes in the mammalian body have been shown to be<br />

proportional to the body weight <strong>of</strong> the animal (Adolph, 1949; Calabrese,<br />

1983; Peters, 1983; Chappell and Mordenti, 1991) and can be related by<br />

allometry, a word from the Greek meaning the measurement (metry) <strong>of</strong><br />

changing size (allo). It has been shown that blood flow, organ size,<br />

metabolic and respiratory rate, and many other physiological and<br />

anatomical variables are related by the general allometric equation<br />

(Boxenbaum, 1982b):<br />

(4.1)<br />

where Y is the function to be measured, W the body weight <strong>of</strong> the animal,<br />

a the coefficient and b the exponent. For mammals, whilst a is different for<br />

each function, b is approximately 0.6–0.8 for rates, flows and clearances, 1.<br />

0 for volumes and organ sizes, and 0.25 for cycles and times. Thus<br />

metabolic rate can be calculated from 7.0·W 0.75 , liver blood flow from<br />

37·W 0.85 , blood weight from 0.055·W 0.99 , and respiratory rate from 0.<br />

019·W 0.26 . Since the blood flows and the weights <strong>of</strong> the liver and kidney,<br />

the two major organs <strong>of</strong> elimination, can be similarly allometrically scaled,<br />

it follows that the same formula could in principle be used for<br />

extrapolation <strong>of</strong> the clearance <strong>of</strong> chemicals between species.<br />

In the past there has been much discussion on the possibility <strong>of</strong><br />

predicting human kinetics and distribution from animal data, using<br />

allometry. For industrial chemicals relatively complex physiological models<br />

have been constructed using this knowledge <strong>of</strong> relative blood flows and<br />

organ size to predict what levels <strong>of</strong> exposure could be expected in man<br />

(Andersen et al., 1984), but little work has been published on comparative<br />

interspecies clearances which will dictate the circulating levels. For drugs,<br />

on the other hand, a number <strong>of</strong> reports have been published on the<br />

rationale for the use <strong>of</strong> allometric scaling <strong>of</strong> kinetics (Dedrick, 1973;<br />

Boxenbaum, 1982b, 1984, 1986; Mordenti, 1985, 1986; Sawada et al.,<br />

1985; Chappell and Mordenti, 1991) but many have been concerned with<br />

its theoretical aspects rather than with its practical use for prediction.<br />

When scaling has been used, the predictions have not always been<br />

accurate, and the method has therefore not had wide usage. This is<br />

unfortunate since the ability to predict what will be the blood levels in man,<br />

without the need to administer the compound, can potentially have many<br />

advantages in drug development and in the safety testing <strong>of</strong> industrial<br />

chemicals where dosing volunteers is <strong>of</strong>ten unacceptable.<br />

Methods<br />


A meta-analysis <strong>of</strong> the papers related to this subject has been made from<br />

those published over the last 20 years. Data before this have largely been<br />

rejected due to the poor design <strong>of</strong> the studies or lack <strong>of</strong> analytical


precision. In the main the data have come from drugs but the same general<br />

considerations would hold for environmental chemicals.<br />

Wherever possible the only compounds included in the analysis have<br />

been those where unbound clearance after systemic administration has been<br />

reported, unless it has been shown that there are little interspecies<br />

differences in protein binding or that absorption is known to be complete<br />

in all the animals. In the past these provisos have not always been met,<br />

leading to incorrect interpretation <strong>of</strong> the data. In most reports the<br />

allometric scaling has used results from at least four species but in some<br />

cases up to 11 have been included. Practically this would involve an<br />

enormous resource and would be difficult when many compounds are<br />

being investigated. For this analysis it has been assumed that only one<br />

species will initially be used and the aim <strong>of</strong> this analysis was to find which<br />

single species would provide the best prediction <strong>of</strong> clearance compared to<br />

that found in man.<br />

Three methods have been used using data, wherever possible, from<br />

mouse, rat, rabbit, dog and monkey (macaques) in a total <strong>of</strong> 60<br />

compounds, with human unbound clearances ranging from 4 to 150 909 ml<br />

min −1 .<br />

Simple allometric equation<br />

Figure 4.1 shows a typical allometric relationship for the clearance <strong>of</strong> the<br />

anticancer drug, fotemustine, showing that equation (4.1) can be made<br />

linear for the determination <strong>of</strong> the variables by logarithmically<br />

transforming the body weight (W) and clearances (CL), as shown in<br />

equation (4.2) where the exponent b can be calculated from the slope <strong>of</strong><br />

the linear regression.<br />

(4.2)<br />

From this analysis <strong>of</strong> all the available papers, where this has been<br />

undertaken with more than four species using data taken from 29<br />

compounds, it was possible to show that the mean exponent (b) is<br />

approximately 0.70±0.15 for unbound clearance, but with a range <strong>of</strong> 0.92–<br />

0.28. This mean value is to be expected since it is comparable to the<br />

exponent for the allometric equation relating physiological rates and<br />

clearances to weight as for metabolic rate, body surface area, hepatic and<br />

renal blood flow, etc. Therefore it would seem that even without a specific<br />

knowledge <strong>of</strong> the clearance in a number <strong>of</strong> different species, it could be<br />

assumed that the exponent <strong>of</strong> 0.7 is a common factor for all chemicals, if it<br />

has not been previously determined. The coefficient a can subsequently be<br />

determined for each compound from only one species according to<br />

equation (4.1), and a predictive value for man determined.

Body surface area (BSA)<br />

It has been suggested that the body surface area provides a good measure<br />

<strong>of</strong> overall metabolic rate and that this may be a better measure <strong>of</strong> relative<br />

clearance between species (Chiou and Hsu, 1988). The BSA has therefore<br />

been cal culated for each species using Meehs Formula, BSA=0.103·W 0.67<br />

(Spector, 1956) and the ratio <strong>of</strong> human BSA to animal BSA multiplied by<br />

the animal clearance, to determine the predicted human clearance.<br />

Life span correction<br />


Figure 4.1. Allometric scaling <strong>of</strong> Fotemustine clearance compared with the body<br />

weight in various species.<br />

For some drugs, particularly those which are extensively metabolised but<br />

have a low hepatic clearance, such as phenytoin, antipyrine or caffeine<br />

(Boxenbaum, 1982b; Bonati et al., 1984–5), these simple scaling methods<br />

seem to be poorly predictive for man and an allometric correction using<br />

maximum life potential (MLP) has been used to improve the accuracy<br />

(Figure 4.2). Although the allometric approach using body weight alone is


Figure 4.2. Comparison <strong>of</strong> the allometric interspecies scaling for phencyclidine<br />

using: (top) clearance (CL), and (bottom) clearance corrected for maximum life<br />

potential (MLP) in seven species (redrawn from Owens et al., 1987).<br />

valid for many physiological functions it is poorly predictive <strong>of</strong> longevity<br />

or maximum life potential in man. Using a derived equation based on body<br />

weight alone, humans should only live for 26.6 years, clearly an<br />

underestimate. In fact Sacher (1959) has shown that a better measurement<br />

<strong>of</strong> life span can be calculated using not only body weight but also brain<br />

weight (equation (4.3)), and with this correction the MLP for man<br />

becomes 113 years (Boxenbaum and De Souza, 1988).

(4.3)<br />

Simplistically it has been suggested that these differences in longevity can<br />

be explained by the assumption that in any one species there is a<br />

predetermined or fixed amount <strong>of</strong> total ‘body metabolic potential’ and<br />

once this is used up the animal dies (Boddington, 1978). Boxenbaum (1986)<br />

has extrapolated this concept to include intrinsic hepatic metabolism<br />

suggesting that there is a certain quantity <strong>of</strong> ‘hepatic pharmacokinetic<br />

stuff’ per unit <strong>of</strong> body weight available in a life-time which can be<br />

interrelated by the formula:<br />

(4.4)<br />

where CL is the unbound clearance, and c is a constant for each compound.<br />

Thus, the longer the animal lives, the slower this ‘stuff’ is used up.<br />

Examination <strong>of</strong> the data available from 13 disparate compounds<br />

(Table 4.1), where at least four species have been investigated, shows the<br />

MLP correction has produced good results with an exponent b equal to<br />

unity. Thus this would suggest that the relative clearance between species is<br />

directly proportional to their body weight (W) and MLP, and that animal<br />

(CL (A)) and human clearance (CL (H)) can be simply related according to<br />

equation (4.4).<br />

(4.5)<br />

The maximum life potential (MLP) has been calculated for each animal<br />

from Sacher’s formula (equation (4.3)) (mouse=2.7 y, rat=4.7 y, dog=20 y,<br />

rabbit=8 y, monkey=22 y and human=113 y).<br />

For each drug where the appropriate information was available, the<br />

human clearance has been calculated from each species using the above<br />

approaches and compared with that observed (Table 4.2), and the<br />

percentage prediction measured as:<br />

Results<br />


The data from 60 different compounds were used in this ongoing analysis<br />

and as could be expected more data were available for the rat (n=47)<br />

compared to mouse (n=27) and dog (n=28), rabbit (n=24), or monkeys<br />

(n=17). In four cases, valproic acid, diazepam, ceftizoxime and<br />

theophylline, different results were found and data have been analysed<br />

separately. For two classes <strong>of</strong> drugs, β-lactams and benzodiazepines, data<br />

from a number <strong>of</strong> compounds were available (n=6 and 12, respectively), but<br />

only mean values were used in this analysis to minimise a class <strong>of</strong><br />

compounds bias in the results.


Table 4.1 Comparison <strong>of</strong> exponential values for b with MLP corrected clearance<br />

(CL u ·MLP=aW b )<br />

From Figure 4.3 it can be seen that for most species the use <strong>of</strong> the simple<br />

exponent 0.7 provided the worst prediction, particularly in the mouse and<br />

dog, which overestimated the human clearance by approximately 600 and<br />

400 per cent, respectively. The rat and rabbit (100–150 per cent) were<br />

better but the monkey was best giving a small overestimate (36 per cent). The<br />

body surface area calculation for most animals gave a better result<br />

particularly for the rat (48 per cent) and monkey (−28 per cent), but the<br />

best method overall is the use <strong>of</strong> the maximum life potential correction<br />

which provided reasonable predictions, within 50 per cent, for all species<br />

with the exception <strong>of</strong> the mouse (89 per cent). The mean accuracy values<br />

only provide part <strong>of</strong> the picture on predictions and the variation, range and<br />

outliers can give additional information on precision and confidence <strong>of</strong> the<br />

analyses. Table 4.3 shows that although there is reasonable accuracy with<br />

the rat, rabbit and dog, the coefficients <strong>of</strong> variations and range <strong>of</strong> values<br />

for these species are large, particularly in the dog, even though the mean<br />

value is reasonable. However for the monkey most estimates <strong>of</strong> human<br />

clearance fall within close proximity to the mean provid ing good<br />

confidence in the data. Similarly the number <strong>of</strong> all compounds which have<br />

a predictability <strong>of</strong> more than 100 per cent error was large for the dog (18 per<br />

cent) and mouse (11 per cent), less for the rat and rabbit, but none were<br />

found for the monkey. In the rat, where the largest number <strong>of</strong> compounds<br />

were examined (n=56), there is a good correlation (r=0.81, p

Figure 4.3. Mean prediction values (percentage error) for human clearance<br />

calculated for various species using: exponent 0.7, body surface area (BSA), and<br />

maximum life potential correction (MLP).<br />


For these life span corrections, equation (4.3) has been used to calculate<br />

MLP, but monkeys in captivity, in contract organisations and zoos<br />

(Carmac, 1994), appear to live longer than the calculated 22 years and<br />

ages <strong>of</strong> 35 years are not uncommon. Substituting this longer life span into<br />

the clearance MLP correction improves the mean accuracy to −14 per cent,<br />

but the range increases and 2 per cent <strong>of</strong> compounds now give a prediction<br />

greater than 100 per cent. Attempts to combine predictions from two or<br />

more animals did not improve the accuracy <strong>of</strong> the predictions but did<br />

marginally improve the confidence <strong>of</strong> these values, particularly when the<br />

data from rat and monkey were averaged, from a confidence interval <strong>of</strong><br />

±20 and ±23 for rat and monkey respectively, when used alone, to ±15<br />

when the results were combined.<br />

From this analysis <strong>of</strong> the data it would appear that measurement <strong>of</strong> the<br />

clearance <strong>of</strong> a drug in the monkey together with a correction for MLP<br />

differences, provide the best overall estimate <strong>of</strong> human clearance with the<br />

greatest confidence in the results, although for many compounds the rat or<br />

even the rabbit are good alternatives. The mouse and the dog, on the other<br />

hand, seem to be poorer animal models to extrapolate to human kinetics.<br />

Discussion<br />

There has in the past been a hesitation to use allometric scaling to predict<br />

the clearance in man, but it would appear from this review <strong>of</strong> the literature<br />

that this approach can be used for predictive purposes with an acceptable<br />

degree <strong>of</strong> accuracy, even when the clearance is measured in only one<br />

species. To put this in perspective, if the actual human clearance was 500 ml<br />

min−1 , the predicted clearance using rat or monkey with MLP correction<br />

would be a oximately 300 ml min 1 ppr<br />

− with a 95 per cent confidence,


Table 4.2 Human unbound clearances <strong>of</strong> the compounds used in this analysis


a Campbell DB, 1993 unpublished data.<br />

b CL=812 ml min−1


Table 4.3 Interspecies comparisons <strong>of</strong> human clearance predictions expressed as<br />

percentage from observed clearances using maximum life potential corrections<br />

a MLP=22 years.<br />

b MLP=35 years.<br />

c Percentage <strong>of</strong> compounds with a predicted human clearance greater than 100% <strong>of</strong><br />

that observed.<br />

Figure 4.4. Relationship between observed human clearance and that calculated<br />

from the rat using a maximum life potential (MLP) correction (n=56) (—— line <strong>of</strong><br />

identity).<br />

ranging from 240 to 360 ml min− 1.<br />

The monkey appears to be slightly<br />

better than the rat and rabbit in terms <strong>of</strong> the accuracy <strong>of</strong> prediction, and<br />

with a few exceptions may be phylogenetically more acceptable. This is<br />

perhaps not surprising since most studies which have examined species<br />

differences in metabolism indicate that the monkey is more similar to man<br />

compared to the rat (Caldwell, 1981). All the primate data reported, as far<br />

as can be ascertained, have come from the Rhesus or Cynomolgus, Old<br />

World Macaque monkeys. The same considerations may not be true for<br />

New World monkeys, such as the squirrel or marmoset, but few kinetic<br />

comparisons have been made with these species.<br />

In practice, prediction <strong>of</strong> human clearance would involve measuring the<br />

intravenous or intramuscular kinetics, namely the infinite area under the<br />

curve, for each investigatory compound in two to four animals, together<br />

with an estimate <strong>of</strong> the in vitro protein binding in the animal under<br />

investigation and in human plasma, to obtain the free intrinsic clearance

and then multiply the animal clearance by the ratio <strong>of</strong> weight and MLP,<br />

approximately 13 for the rat and 3.5 for a macaque monkey. Of course, as<br />

shown by these data, there can be exceptions, and the monkey and indeed<br />

the rat may not be a suitable species to undertake allometric scaling for all<br />

compounds. However there is an increasing use <strong>of</strong> in vitro systems such as<br />

isolated microsomes, hepatocytes or hepatic slices, to compare the<br />

metabolic pr<strong>of</strong>iles <strong>of</strong> compounds in animals. If undertaken in conjunction<br />

with allometric scaling, pr<strong>of</strong>ound interspecies differences in the rates and<br />

extent <strong>of</strong> metabolism compared to humans could be observed and provide<br />

information on which is the most suitable species to use for scaling. Since<br />

the allometric scaling for volume appears for most compounds to be<br />

directly proportional to body weight with an exponent <strong>of</strong> approximately 1.<br />

0, half-life can also be easily calculated thereby providing all the necessary<br />

kinetic parameters to simulate plasma levels after repeated dosing in man.<br />

With this information the absolute need to undertake kinetic analysis <strong>of</strong><br />

industrial chemicals in volunteers would be reduced since the exposure<br />

calculated by this procedure is considerably better than that employed<br />

presently using uncertainty factors, giving errors in excess <strong>of</strong> 1000 per<br />

cent.<br />

Further studies are <strong>of</strong> course needed to confirm these initial<br />

observations, particularly with those chemicals used in industry or potential<br />

environmental pollutants, but perhaps this re-evaluation shows that<br />

allometry, when correctly used, may well have a practical role in the<br />

evaluation <strong>of</strong> their potential risk to man.<br />

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PART TWO<br />

Reactive industrial chemicals

5<br />

Metabolism <strong>of</strong> Reactive Chemicals<br />

PETER J.van BLADEREN 1,2 and BEN van OMMEN 1<br />

1 TNO <strong>Toxicology</strong>, Zeist<br />

2 Agricultural University, Wageningen<br />

Introduction<br />

For the purpose <strong>of</strong> the present paper, a reactive chemical will be defined as<br />

a strongly electrophilic agent. Such compounds can bind to the numerous<br />

macromolecular targets in the cell, and thus elicit toxic effects. Binding to<br />

DNA can result in mutations or cancer, binding to proteins or membrane<br />

components to cytotoxicity or specific forms <strong>of</strong> toxicity.<br />

A scale could be drawn up for the reactivity <strong>of</strong> electrophiles. However, it<br />

is not certain that those compounds on the high end <strong>of</strong> the scale, i.e. the<br />

most reactive, would also be the most toxic. On the contrary, these<br />

compounds might be expected to react quickly with water and thus not<br />

reach their target molecules. For the purpose <strong>of</strong> classifying the reactivity <strong>of</strong><br />

electrophiles, the most useful is the theory <strong>of</strong> s<strong>of</strong>t and hard acids and bases<br />

(e.g. Commandeur and Vermeulen, 1990). In principle, the preferential<br />

targets for electrophiles can be derived. Furthermore, an electrophile<br />

showing the highest affinity for the relatively hard nitrogen and oxygen<br />

nucleophiles <strong>of</strong> DNA may pose a higher risk for mutations and cancer than<br />

one reacting preferentially with s<strong>of</strong>t sulfur nucleophiles such as found in<br />

proteins and glutathione.<br />

The following classes <strong>of</strong> electrophiles will be discussed: (1) quinones,<br />

which can both arylate as well as cause toxicity through redox cycling; (2)<br />

derivatives with an actual leaving group such as methylene chloride and<br />

ethylene dibromide, and (3) reagents such as isothiocyanates, isocyanates<br />

and α,β-unsatu-rated<br />

ketones and aldehydes.<br />

The enzymes involved in activation and detoxication<br />

To become toxic, almost all <strong>of</strong> the chemicals to which man is exposed,<br />

including the carcinogens, need metabolic activation. The reactive<br />

intermediates that are formed during metabolism are responsible for<br />

binding to cellular macromolecules which very likely elicit the toxic


response. In general, other biotransformation enzymes can detoxify these<br />

metabolites. Thus, the concentration <strong>of</strong> the ultimate carcinogen, or<br />

toxicant in general, is the result <strong>of</strong> a delicate balance between the rate <strong>of</strong><br />

activation and the rate <strong>of</strong> detoxification. Although toxicological processes<br />

can be much more complex, interindividual differences in susceptibility are<br />

certainly also a result <strong>of</strong> interindividual differences in this balance between<br />

metabolic activation and detoxification.<br />

The enzymes which are to a large extent responsible for the formation <strong>of</strong><br />

reactive metabolites belong to the family <strong>of</strong> cytochromes P-450. However,<br />

for almost all enzymes involved in biotransformation, examples have been<br />

described <strong>of</strong> activation <strong>of</strong> specific classes <strong>of</strong> chemicals. The main classes <strong>of</strong><br />

enzymes involved in detoxifying chemicals which are reactive per se as well<br />

as reactive metabolites are the epoxide hydrolases and the glutathione Stransferases.<br />

NADPH quinone reductase is involved in the reduction <strong>of</strong><br />

quinones.<br />

Epoxide hydrolases<br />

Metabolites which contain an epoxide moiety may undergo hydrolytic<br />

cleavage to less reactive vicinal dihydrodiols. This reaction is catalyzed by<br />

the enzyme epoxide hydrolase (EH), which was first thought to be<br />

exclusively located in the endoplasmic reticulum (microsomal epoxide<br />

hydrolase, mEH; Oesch, 1972). In later studies on the mammalian<br />

metabolism <strong>of</strong> certain alkyl epoxides, the existence <strong>of</strong> a cytosolic EH (cEH)<br />

was demonstrated (Gill et al., 1974). The two forms <strong>of</strong> EH have<br />

complementary substrate specificity, in that many epoxides, e.g. arene<br />

oxides, which are good substrates for mEH are poor substrates for cEH,<br />

and vice versa, e.g. trans-disubstituted oxiranes are good substrates for cEH<br />

but not for mEH (Hammock and Hasagawa, 1983). Other studies have<br />

pointed to the fact that the common nomenclature <strong>of</strong> ‘microsomal’ and<br />

‘cytosolic’ epoxide hydrolase is not semantically precise: metabolic and<br />

immunochemical studies demonstrated the existence <strong>of</strong> membrane-bound<br />

forms <strong>of</strong> cEH (Guenthner and Oesch, 1983), whereas mEH-like activity<br />

was detected in cytosolic fractions <strong>of</strong> human tissue (Schladt et al., 1988).<br />

Glutathione S-transferases<br />

Glutathione is involved in a variety <strong>of</strong> vital cellular reactions. First, a large<br />

number <strong>of</strong> the various classes <strong>of</strong> xenobiotics to which man is exposed—<br />

industrial, therapeutic as well as naturally occurring chemicals—are<br />

metabolized in vivo to reactive intermediates. Such electrophilic<br />

metabolites may bind to cellular macromolecules and thus cause toxicity.<br />

The formation <strong>of</strong> glutathione conjugates, both by spontaneous reaction<br />

between the reactive species and glutathione as well as catalyzed by the


glutathione S-transferases, is the main detoxification mechanism for<br />

electrophiles in mammalian cells (Chasseaud, 1979). Secondly, via<br />

glutathione peroxidase and the glutathione S-transferases, hydrogen<br />

peroxide and organic peroxides are detoxified, yielding glutathione disulfide<br />

as one <strong>of</strong> the products (Prohaska, 1980). Thirdly, glutathione and the<br />

glutathione S-transferases play a role in the biosynthesis <strong>of</strong> such important<br />

endogenous compounds as prostaglandins and leukotriene C4 (Söderstrom<br />

et al., 1985; Ujihara et al., 1988). In fact, in the latter case one may argue<br />

that an endogenous compound is activated by conjugation with<br />

glutathione, since leukotriene C4 is a mediator <strong>of</strong> the adverse reactions<br />

associated with asthmatic attacks (Samuelson, 1988).<br />

The GSTs are a family <strong>of</strong> isoenzymes with broad and overlapping<br />

substrate selectivity. Although membrane-bound forms <strong>of</strong> GST have been<br />

detected (Morgenstern et al., 1988), GST activity is mainly located in the<br />

cytosol. GSTs are dimers <strong>of</strong> subunits and within a dimer, each subunit<br />

functions independently <strong>of</strong> the other (Mannervik and Jensson, 1982). The<br />

GSTs are now known to be a multi-gene family <strong>of</strong> isoenzymes, which can<br />

be divided into four classes (alpha, mu, pi and theta), based on similarity in<br />

structural, physical and catalytic properties <strong>of</strong> their subunits (Ketterer and<br />

Mulder, 1990; Vos and Van Bladeren, 1990). In addition to their crucial role<br />

in catalyzing glutathione conjugation, GSTs may also be important in<br />

intracellular binding and/or transport <strong>of</strong> endogenous and xenobiotic nonsubstrate<br />

ligands (Listowsky et al., 1988).<br />

The glutathione conjugates initially formed from electrophilic species are<br />

further processed via -glutamyltranspeptidase which splits <strong>of</strong>f the<br />

glutamate residue, and dipeptidases which remove the glycine moiety. The<br />

resultant cysteine S-conjugates are then acetylated to give so-called<br />

mercapturic acids which are excreted into the urine (Jakoby, 1980).<br />

Interestingly, mercapturic acids were the first metabolites derived from<br />

xenobiotics to be recognized as such (Baumann and Preusse, 1879).<br />

In recent years it has become increasingly evident that glutathione<br />

conjugation is also involved in the formation <strong>of</strong> toxic metabolites from a<br />

variety <strong>of</strong> chemicals (Monks et al., 1990b). These metabolites display a<br />

wide spectrum <strong>of</strong> toxic effects, ranging from cytotoxicity to genotoxicity.<br />

The various mechanisms elucidated for the toxic action <strong>of</strong> the conjugates<br />

can be grouped as follows: (1) directly toxic glutathione conjugates may be<br />

formed from vicinal and geminal dihaloalkanes, via the formation <strong>of</strong> sulfur<br />

halfmustards; (2) from several types <strong>of</strong> glutathione conjugates active<br />

metabolites may be formed by further metabolic steps: conjugates <strong>of</strong><br />

hydroquinones can be oxidized to give reactive quinones, and conjugates<br />

derived from haloalkenes are transformed into electrophilic species by the<br />

action <strong>of</strong> cysteine conjugate β-lyase. For both hydroquinones and<br />

haloalkenes the selective nephrotoxicity observed is the result <strong>of</strong> the<br />

targeting <strong>of</strong> the conjugates to the kidneys; (3) glutathione conjugates may


serve as transporting and targeting agents for compounds that react<br />

reversibly with gluathione such as isothiocyanates, isocyanates and α,<br />

βunsaturated<br />

ketones (Van Bladeren, 1988).<br />

Glutathione S-transferase polymorphism<br />

Genetic variation in the expression <strong>of</strong> GST isoenzymes has been studied<br />

almost solely in man. Considerable variation, possibly indicating a<br />

polymorphism, has been observed for the human liver alpha class<br />

isoenzymes. The ratio <strong>of</strong> GSTA1 and GSTA2 subunits, as determined by<br />

HPLC, was found to range from 0.5 to over 10 (Van Ommen et al., 1990).<br />

However, a division into two groups, with average ratios <strong>of</strong> 1.6±0.3 and 3.<br />

8±0.6 could be made, suggesting an alpha class polymorphism. In view <strong>of</strong><br />

the fact that subunits GSTA1 and GSTA2 together make up a major<br />

portion <strong>of</strong> the GST protein in human liver this potential polymorphism<br />

merits further attention. For class mu isoenzymes a clear polymorphism<br />

has been observed in humans: iso-enzyme GSTM1a-1a was found to be<br />

expressed in only 60% <strong>of</strong> the samples analyzed (Board, 1981). In this study<br />

no account was taken <strong>of</strong> the fact that a second mu class isoenzyme,<br />

isoenzyme GSTM1b-1b was also suggested to play a part in this<br />

polymorphism. In a study on the excretion <strong>of</strong> the mercapturate derived<br />

from 1,3-dichloropropene in exposed workers, however, no difference was<br />

observed between mu-positive and mu negative subjects (Vos et al., 1991).<br />

Quinones and their glutathione conjugates<br />

Two modes <strong>of</strong> reactivity can form the basis <strong>of</strong> the toxicity associated with<br />

quinones: (i) their ability to undergo ‘redox cycling’ and to thereby create<br />

an oxidative stress (Kulkarni et al., 1978), and (ii) their electrophilicity<br />

allowing them to react directly with cellular nucleophiles such as protein<br />

and non-protein sulfhydryls (Dierickx, 1983). Since glutathione is the<br />

major non-protein sulfhydryl present in cells, it comes as no surprise that it<br />

is intimately involved in the biological effects <strong>of</strong> quinones. On the one<br />

hand, glutathione can act as a reducing agent, detoxifying quinones by<br />

converting them to hydroquinones with the concomitant formation <strong>of</strong><br />

glutathione disulfide. On the other hand quinone and hydroquinonethioethers<br />

are formed. Recently considerable evidence has been gathered,<br />

indicating that a variety <strong>of</strong> these thioethers possess biological activity<br />

(Dierickx, 1983; Koga et al., 1986).<br />

The target sites for the biological (toxicological) activity <strong>of</strong><br />

quinonethioethers is to a large extent determined by the glutathione moiety:<br />

as will be discussed, the main targets are the kidney (Monks et al., 1985)<br />

and various enzymes using glutathione as a (second) substrate, e.g. the<br />

glutathione S-transferases (Van Ommen et al., 1988). Bromobenzene is


toxic to proximal renal tubules. The nephrotoxic effect <strong>of</strong> o-bromophenol<br />

and bromo hydroquinone was found to be considerably higher, indicating<br />

that these compounds were situated along the main bioactivation route<br />

(Monks et al., 1985). Subsequent elegant work by Monks and Lau has<br />

shown that in fact the nephrotoxicity is caused by the glutathione<br />

derivatives <strong>of</strong> bromohydroquinone (Lau and Monks, 1990). Interestingly,<br />

the relative toxicity <strong>of</strong> the quinoneglutathione conjugates increases as the<br />

extent <strong>of</strong> glutathione addition increases, i.e. the diglutathionyl derivative is<br />

more toxic than the monoconjugate (Monks et al., 1988b). The tissue<br />

selectivity is a consequence <strong>of</strong> their targeting to renal proximal tubule cells<br />

by the brushborder -glutamyl transpeptidase. AT-125, a selective inhibitor<br />

<strong>of</strong> this enzyme in vivo, protects the kidney from the toxic effects <strong>of</strong> the<br />

conjugates. The toxicity <strong>of</strong> these hydroquinone conjugates is apparently<br />

not mediated by cysteine conjugate β-lyase catalyzed formation <strong>of</strong> thiols.<br />

The inhibitor <strong>of</strong> the lyase, amino-oxyacetic acid, had only minor effects on<br />

the extent <strong>of</strong> toxicity, and the putative product, 6-bromo-2,5dihydroxythiophenol,<br />

needed activation by oxidation before it exerted any<br />

biological effect (Monks et al., 1990b). Thus, the effects <strong>of</strong> these<br />

conjugates apparently are a consequence <strong>of</strong> their oxidation to the<br />

corresponding quinones.<br />

Several isomers <strong>of</strong> 2-bromo-glutathionyl as well as the<br />

bromodiglutathionyl hydroquinones were isolated and tested. Instead <strong>of</strong> a<br />

direct correlation <strong>of</strong> toxicity with the electrochemical properties <strong>of</strong> these<br />

compounds, it was found that the diglutathionyl derivative, which is by far<br />

the most toxic, was the most stable to oxidation at pH 7.4 (Monks and<br />

Lau, 1990). The paradox was clarified by Monks and Lau by determining<br />

the oxidation potentials <strong>of</strong> the breakdown products for the mercapturic<br />

acid pathway: hydrolysis <strong>of</strong> the glutathione moiety gives rise to the cysteine<br />

derivative, which is more readily oxidized than the parent compound<br />

(Monks and Lau, 1990). Apparently two detoxication pathways are<br />

possible for these cysteine derivatives: N-acetylation results in formation <strong>of</strong><br />

the mercapturic acid which again is relatively resistant to oxidation, but<br />

oxidative cyclization <strong>of</strong> cysteinylglycine and cysteine derivatives has been<br />

found to give 1,4-benzothiazines, which do not possess any apparent toxic<br />

properties (Monks and Lau, 1990). The action <strong>of</strong> -glutamyl<br />

transpeptidase can thus result in both activation as found for 2bromohydroquinone<br />

derivatives, but also in detoxication as was observed<br />

for 2,5-dichloro-3-(glutathion-S-yl)hydroquinone and 2,5,6-trichloro-3glutathion-S-yl)hydroquinone<br />

(Mertens et al., 1991). The ease with which<br />

the 1,4-benzothiazines are formed is very likely the determining factor in this<br />

case. A similar pathway has been worked out for p-aminophenol, a known<br />

nephrotoxic metabolite <strong>of</strong> acetaminophen (Eckert et al., 1989, 1990).<br />

Bioactivation <strong>of</strong> halogenated benzenes has long been thought to be the<br />

result <strong>of</strong> oxidation to an epoxide. However, recent studies have shown that


the covalent binding to cellular macromolecules is not only the result <strong>of</strong> the<br />

first oxidative step, but also <strong>of</strong> the second, the formation <strong>of</strong> a quinone or<br />

hydroquinone from the initially formed phenol. The quinone in turn can be<br />

detoxified by glutathione conjugation. However, although glutathione<br />

protects the liver against toxicity due to these quinones, the conjugates are<br />

transported to the kidney and are there activated to new reactive<br />

intermediates. Thus, increasing the relative amount <strong>of</strong> glutathione Stransferases<br />

in this case would not really protect the organism, but merely<br />

change the target organ <strong>of</strong> the active metabolites.<br />

Chemicals with a leaving group<br />

Methylene chloride<br />

Both vicinal and geminal haloalkanes are bioactivated via conjugation with<br />

glutathione. The glutathione-dependent metabolism <strong>of</strong> the important<br />

industrial solvent dichloromethane yields S-chloromethyl-glutathione as the<br />

initial metabolite (Ahmed and Anders, 1976). This intermediate is held<br />

responsible for the carcinogenicity <strong>of</strong> dichloromethane in the mouse.<br />

Interestingly, this compound does not cause tumors in rats, and this has<br />

been related to the fact that the rate <strong>of</strong> metabolism via the glutathione<br />

pathway, catalyzed by the glutathione S-transferases, is much lower in rat<br />

tissue than in mouse tissue. Man has been postulated to resemble the rat in<br />

this respect and is thus presumably safe from the carcinogenic effects <strong>of</strong><br />

methylene chloride (ECETOC, 1988). When it does not react with cellular<br />

macromolecules, the intermediate S-chloromethyl-glutathione is converted<br />

non-enzymatically to S-hydroxymethyl-glutathione, which easily eliminates<br />

formaldehyde and regenerates glutathione (Ahmed and Anders, 1978).<br />

The glutathione S-transferase isoenzyme involved in the formation <strong>of</strong> Schloromethylglutathione<br />

belongs to class theta. Interestingly, a<br />

considerable amount <strong>of</strong> interindividual variation could be observed in a<br />

group <strong>of</strong> 22 individuals (Bogaards et al., 1993).<br />

1,2-Dibromoethane and 1,2-dichloroethane<br />

The vicinal dihaloalkanes are exemplified by 1,2-dibromoethane and 1,2dichloroethane,<br />

which are mutagenic, carcinogenic as well as nephrotoxic<br />

(Van Bladeren et al., 1980; Wong et al. 1982; Guengerich et al., 1984;<br />

Elfarra and Anders, 1985; Cheever et al., 1990). The metabolism <strong>of</strong> these<br />

compounds involves two pathways, cytochrome P-450 dependent<br />

oxidation and glutathione S-transferase catalyzed formation <strong>of</strong> glutathione<br />

conjugates. The oxidative pathway results in chloro- and<br />

bromoacetaldehyde, respectively. These aldehydes are electrophilic and


thought to be responsible for the covalent binding <strong>of</strong> 1,2-dichloro- and 1,2dibromoethane<br />

metabolites to protein (Inskeep and Guengerich, 1984).<br />

Glutathione conjugation results in the formation <strong>of</strong> S-2haloethylglutathione<br />

derivatives, which are sulfur halfmustards and as such<br />

highly reactive metabolites (Jean and Reed, 1989). The formation <strong>of</strong> these<br />

conjugates is catalyzed by the glutathione S-transferases, and both in rat<br />

and man the alpha-class isoenzymes have been found to be the most<br />

efficient in this catalysis (Cmarik et al., 1990).<br />

The glutathione pathway is responsible for the mutagenicity (Van<br />

Bladeren et al., 1980), the DNA-binding (Koga et al., 1986) as well as very<br />

likely the carcinogenicity (Cheever et al., 1990) <strong>of</strong> 1,2-dichloro- and 1,2dibromoethane.<br />

The S-2-haloethylglutathione derivatives are strong<br />

alkylating agents (e.g. Jean and Reed, 1989). Their electrophilicity is<br />

attributable to neighboring-group assistance. The halogen atom is displaced<br />

by the sulfur atom on the next carbon atom, to form a highly reactive<br />

episulfonium ion. The intermediacy <strong>of</strong> this reactive species is supported by<br />

stereochemical studies as well as NMR data (Van Bladeren et al., 1979;<br />

Dohn and Casida, 1987; Peterson et al., 1988).<br />

The relative importance <strong>of</strong> the oxidative and glutathione-dependent<br />

pathway in vivo is difficult to determine, since both pathways give rise to<br />

the formation <strong>of</strong> the same 2-hydroxyethylmercapturate. Using<br />

tetradeutero-1,2-dibromoethane, the ratio <strong>of</strong> the pathways has been<br />

calculated as 4:1 (Van Bladeren et al., 1981b). However, isotope effects<br />

might have a considerable influence on this ratio (White et al., 1983).<br />

The major DNA-adduct derived from 1,2-dibromoethane has been<br />

identified by Guengerich and coworkers to be S-(2-(N7-guanyl)-ethyl)<br />

glutathione (Ozawa and Guengerich, 1983; Koga et al., 1986). In addition,<br />

the structure <strong>of</strong> one <strong>of</strong> several minor adducts was recently found to be S-(2-<br />

(Nl-adenyl)ethyl) glutathione (Dong-Hyun et al., 1990). A series <strong>of</strong> S-2haloethylglutathione<br />

and -cysteine derivatives has been synthesized: all<br />

were found to react with DNA, specifically with guanine residues. As<br />

expected for a mechanism known to involve an intermediate episulfonium<br />

ion, adduct levels were similar for chloro- and bromo-substituted<br />

derivatives. However, in Salmonella typhimurium TA100 a large variation<br />

was observed in the ratio <strong>of</strong> mutations <strong>of</strong> adducts, indicating that the<br />

structure <strong>of</strong> the adduct has a major influence on the mutagenicity<br />

(Humphreys et al., 1990).<br />

Not all vicinal dihaloalkanes seem to give rise to the formation <strong>of</strong><br />

episulfonium ions. Methyl substitution for instance effectively hinders the<br />

mutagenicity through this pathway (Van Bladeren, et al., 1981a) and<br />

studies on 1,2-dibromopropane (Zoetemelk et al., 1986) and hexadeuterol,2-dichloro-propane<br />

(Bartels and Timchalk, 1990) indicate that the<br />

resulting mercapturic acids are only formed through an oxidative pathway.<br />

However, for the heavily used agricultural chemical l,2-dibromo-3-


chloropropane evidence has been accumulating recently, implicating a<br />

glutathione-mediated activation pathway in the renal and testicular toxicity<br />

associated with this compound (Pearson et al., 1990). Interestingly,<br />

consecutive formation <strong>of</strong> two episulfonium ions can occur, and in fact bis-<br />

DNA-adducts have been identified (Humphreys et al., 1991). 1,2-<br />

Dibromochloropropane could thus cross-link DNA strands as the initial<br />

step leading to cell death.<br />

Isoenzyme selectivity for both primary reactions has been studied<br />

extensively. The alpha and theta class glutathione S-transferases are<br />

responsible for the conjugation <strong>of</strong> EDB both in rats and man. For both <strong>of</strong><br />

these enzymes enormous differences in levels between individuals have been<br />

found, which may be due to genetic differences, but are certainly also<br />

influenced by induction. One might expect individuals with an increased<br />

relative amount <strong>of</strong> glutathione S-transferases to be at increased risk.<br />

Reversible glutathione conjugates acting as transporting<br />

agents<br />

Numerous substrates for glutathione conjugation exist where a formal<br />

addition takes place: both the glutathionyl residue and the hydrogen atom<br />

are added to the acceptor molecule. From a chemical point <strong>of</strong> view, this<br />

reaction should be relatively easily reversible. Of course, the extent <strong>of</strong> the<br />

occurrence <strong>of</strong> the reverse reaction depends on the position <strong>of</strong> the<br />

equilibrium and is influenced by such conditions as the concentration <strong>of</strong><br />

the reactants and the pH. The biological consequences <strong>of</strong> this reaction<br />

sequence would be that the original electrophile is detoxified initially, but<br />

not permanently: it can be released again and thus appear in unexpected<br />

parts <strong>of</strong> the body. The glutathione conjugate serves as a storage or<br />

transport form for the alkylating agent. Systemic effects <strong>of</strong> highly reactive<br />

compounds might be explained in this way.<br />

For both isothiocyanates and isocyanates evidence for this pathway has<br />

been obtained. Benzyl and allyl isothiocyanate are both naturally occurring<br />

compounds that are excreted mainly as mercapturic acids in urine after<br />

administration to rats (Brüsewitz et al., 1977). However, the mercapturate<br />

in urine is unstable under basic conditions and reforms the free<br />

isothiocyanate. The glutathione, cysteine as well as N-acetyl-cysteine<br />

conjugates derived from these isothiocyanates are all toxic in vitro<br />

(Bruggeman et al., 1986, Temmink et al., 1986). In vivo, the fact that the<br />

conjugates are somewhat more unstable in urine probably plays a role in<br />

the effects. Benzyl isothiocyanate is used for the treatment <strong>of</strong> bladder<br />

infections (Brüsewitz et al., 1977), while allyl iso-thiocyanate causes<br />

bladder tumors in male rats (Dunnick et al., 1982).<br />

The extremely reactive and toxic methyl isocyanate, used in the<br />

manufacture <strong>of</strong> carbamate pesticides, was released into the atmosphere in


large amounts during a disaster in 1984. To explain the systemic effects <strong>of</strong><br />

exposure to this compound, Baillie and coworkers hypothesized that these<br />

are mediated by the glutathione conjugates (Pearson et al., 1990). In fact, a<br />

rapid distribution <strong>of</strong> radioactivity throughout the body was found for rats<br />

exposed to 14C-methyl isocyanate vapor (Ferguson et al., 1988), the<br />

glutathione conjugate was identified in bile (Pearson et al, 1990) and the<br />

mercapturic acid was identified as a major urinary metabolite (Slatter et<br />

al., 1991) <strong>of</strong> rats dosed with methyl isocyanate. As was found for the<br />

isothiocyanates, in aqueous solution the synthetic glutathione conjugates<br />

are in equilibrium with the free electrophiles and glutathione: when an<br />

excess <strong>of</strong> cysteine is added to the solution, the corresponding cysteine<br />

conjugate is formed rapidly (Pearson et al., 1990). It should be realized<br />

however, that although thiols are the prime targets <strong>of</strong> iso- thiocyanates and<br />

isocyanates, the reactions with oxygen and nitrogen nucleophiles also<br />

occur and give rise to adducts that are much more stable (Pearson et al.,<br />

1991).<br />

The veterinary drug furazolidone is metabolized to a reactive metabolite<br />

that possesses an α,<br />

β-unsaturated ketone functionality. A reversible, socalled<br />

Michael adduct <strong>of</strong> this metabolite with glutathione was identified<br />

and has been suggested to play a role in the toxic effects <strong>of</strong> furazolidone<br />

(Vroomen et al., 1987). In fact residues <strong>of</strong> this metabolite covalently bound<br />

to microsomal protein could be trapped by an excess <strong>of</strong> mercaptoethanol<br />

and the glutathione conjugate gives rise to covalent binding to microsomal<br />

protein (Vroomen et al., 1988). Similarly, 2-methylfuran is metabolized to<br />

acetyl acrolein. The glutathione conjugate derived from this metabolite is<br />

unstable, and in fact toxicity <strong>of</strong> 2-methylfuran is potentiated by increasing<br />

glutathione levels by the administration <strong>of</strong> the cysteine precursor L-2oxothiazolidine-4-carboxylate<br />

(Ravindranath and Boyd, 1991).<br />

Thus, the reversibility <strong>of</strong> glutathione conjugation reactions warrants<br />

further investigation. The fact that reactive intermediates can be reformed<br />

might have important implications for the explanation <strong>of</strong> effects at sites<br />

distant from the site <strong>of</strong> initial exposure and/or initial conjugation.<br />

Conclusion<br />

Reactive chemicals can be detoxified fairly efficiently by several ubiquitous<br />

biotransformation enzymes. However, numerous cases have been reported<br />

where the initial detoxification is not the end <strong>of</strong> the story. The various<br />

pathways that the initially formed metabolites may undergo can result in<br />

unexpected toxicities at sites distant from the point <strong>of</strong> entry into the body<br />

<strong>of</strong> the electrophilic xenobiotic or the site <strong>of</strong> formation <strong>of</strong> the electrophilic<br />



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6<br />

Methods for the Determination <strong>of</strong> Reactive<br />

<strong>Compounds</strong><br />


CIBA-GEIGY Ltd, Basel<br />

Introduction<br />

It has well been recognised for a long time that adverse effects <strong>of</strong> chemicals<br />

are associated with their reactivity whereby many unreactive chemicals are<br />

metabolised in the cell to a reactive intermediate. Reactive intermediates<br />

are generally electrophiles which undergo reactions with cellular<br />

nucleophiles and the toxicological response is <strong>of</strong>ten the consequence <strong>of</strong> the<br />

covalent binding <strong>of</strong> a chemical to cellular macromolecules. This chapter<br />

will provide a brief survey <strong>of</strong> current methods for the determination <strong>of</strong><br />

protein and DNA adducts generated by reactive compounds, and will<br />

discuss some useful applications <strong>of</strong> these technologies.<br />

Source <strong>of</strong> reactive metabolites<br />

Reactive chemicals are generally strong electrophilic agents. These<br />

compounds can be reactive per se (direct electrophiles), such as methylmethanesulphonate,<br />

epoxides or strained lactones. On the other hand,<br />

unreactive chemicals can be enzymatically converted to electrophilic agents<br />

(indirect electrophiles), such as aromatic amines and nitroarenes to the<br />

corresponding nitrenium ions, polycyclic aromatic hydrocarbons to diol<br />

epoxides or N-nitroso compounds to carbenium ions (Figure 6.1; Magee et<br />

al., 1975; Weissburger and Williams, 1975; Lutz, 1979).<br />

Interaction <strong>of</strong> reactive compounds with cellular<br />

constituents<br />

As electrophiles, these reactive compounds undergo reactions with<br />

nucleophiles. The nature <strong>of</strong> the toxicological response is dependent on the<br />

biological macromolecule affected. Reactions with water or glutathione,<br />

two <strong>of</strong> the most abundant cellular nucleophiles, in most cases, lead to an<br />

inactivation <strong>of</strong> the respective reactive compound.


Figure 6.1 Examples <strong>of</strong> electrophilic compounds. The arrows indicate the suspected<br />

electrophilic centres (from Lutz, 1979).


Figure 6.2 Nucleophilic centres in nucleobases and DNA ‘adduct library’ indicating<br />

the preferential binding sites on guanine for several classes <strong>of</strong> chemicals. The most<br />

reactive targets are indicated by an arrow (from Lutz, 1979; Beach and Gupta,<br />

1992).<br />

The most important nucleophilic centres in proteins are the side chains<br />

<strong>of</strong> the amino acids cysteine, methionine, histidine and tyrosine, and the<br />

amino group <strong>of</strong> the N-terminal amino acid. Reactions <strong>of</strong> electrophiles with<br />

proteins lead to the formation <strong>of</strong> protein adducts. This results in general or<br />

specific cytotoxicity depending on whether the function <strong>of</strong> a particular<br />

protein is disturbed (Lawley, 1976; Brooks, 1977). Adduct formation with<br />

blood proteins can result in the formation <strong>of</strong> immunogens and subsequent<br />

allergenic responses.<br />

Finally, reactions with DNA predominantly occur with the nucleobases<br />

adenine, cytosine, thymine and guanine, whereby the most important<br />

nucleophile in DNA is guanine (Figure 6.2). Adduct formation with<br />

nucleobases in DNA is recognised as a crucial step in the formation <strong>of</strong><br />

mutations and cancer (Lutz, 1979).

Methods for the determination <strong>of</strong> adducts<br />

Adduct formation <strong>of</strong> reactive compounds with protein or DNA can easily<br />

be detected by the use <strong>of</strong> radiolabelled test compounds. However, the<br />

radiolabelled compounds are <strong>of</strong>ten not available. In addition, real exposure<br />

situations and unknown mixtures <strong>of</strong> compounds can not be assessed. For a<br />

number <strong>of</strong> chemicals, therefore, alternative methods for adduct<br />

determinations have been developed during the past.<br />

Reactions with proteins can be assessed by analysing haemoglobin or<br />

albumin adducts. These proteins can easily be isolated in large quantities<br />

(100 mg haemoglobin, 30 mg albumin per ml blood) and with sufficient<br />

purity from the blood <strong>of</strong> treated animals or occupationally exposed<br />

humans. Methods for the determination <strong>of</strong> DNA adducts generally require<br />

a higher sensitivity, since DNA from treated animals or exposed humans is<br />

only available in small amounts (1–2 mg per g tissue, 4 µg from white<br />

blood cells per ml blood).<br />

Protein adducts<br />

Physical methods<br />

Aromatic amines and nitroarenes<br />

The key step in the metabolic activation <strong>of</strong> arylamines to the respective<br />

nitrenium ions involves N-hydroxylation. The N-hydroxylamines can be<br />

further oxidised in erythrocytes to the corresponding nitroso compounds<br />

with a concurrent production <strong>of</strong> methaemoglobin. On the other hand,<br />

nitroarenes can be metabolically reduced to the corresponding<br />

nitrosoarenes. The nitrosoarenes covalently bind to the thiol group <strong>of</strong><br />

cysteine residues and rearrange to give stable sulphinic acid amides.<br />

Mild alkaline treatment can be used to hydrolyse these adducts. The<br />

liberated parent amines can be extracted and analysed by HPLC with<br />

specific detection methods, such as electrochemical or fluorescence<br />

detection. In order to improve the sensitivity <strong>of</strong> the assay, the extracted<br />

adducts can be derivatised with electrophores and analysed by GC with<br />

electron capture detection or by GC/MS (Bailey et al., 1990; Skipper and<br />

Tannenbaum, 1990; Sabbioni, 1992, 1994).<br />

Polycyclic aromatic hydrocarbons<br />


Polycyclic aromatic hydrocarbons are oxidised by cytochrome P450 to<br />

epoxides, which are rapidly hydrolysed. However, further oxidation to the


Figure 6.3 Modified Edman degradation <strong>of</strong> alkylated N-terminal valine in<br />

haemoglobin (Törnqvist et al., 1986).<br />

respective diol epoxides results in the formation <strong>of</strong> relatively stable<br />

electrophiles which also alkylate cysteine residues in proteins.<br />

Upon mild acid treatment these adducts are liberated as the respective<br />

tetrols. Similar to adducts from aromatic amines, the tetrols can be<br />

extracted and analysed by HPLC with specific detection methods, or by GC<br />

with electron capture detection or by GC/MS after derivatisation with<br />

electrophores (Shugart and Kao, 1985; Weston et al., 1989; Day et al.,<br />

1990).<br />

Alkylating agents<br />

Adducts <strong>of</strong> alkylating agents with the thiol group <strong>of</strong> cysteine, histidine or<br />

the N-terminal amino acids resist alkaline or acid hydrolysis. To determine<br />

the alkylated amino acids the protein is hydrolysed with 6 N HCl and the<br />

amino acids are separated on a anion exchange column. The fractions<br />

containing the alkylated amino acids are derivatised with electrophores and<br />

analysed by GC/MS (van Sittert et al., 1985; Bailey et al., 1987).

Alternatively, the alkylated N-terminal valine <strong>of</strong> haemoglobin can<br />

selectively be cleaved <strong>of</strong>f by a modified Edman degradation with<br />

pentafluorophenyl isothiocyanate (PFPITC). Since alkylation <strong>of</strong> the amino<br />

group favours the reaction, conditions can be selected to exclusively<br />

liberate alkylated N-terminal amino acids whilst leaving the non-adducted<br />

N-terminal valine intact (Törnqvist et al., 1986). The resulting<br />

pentafluorophenyl thiohydantoine (PFPTH) derivative can be extracted and<br />

quantified by GC/MS (Figure 6.3).<br />

Immunological methods<br />

Immunological methods have been developed for the quantification <strong>of</strong><br />

some adducts <strong>of</strong> aromatic amines, polycyclic aromatic hydrocarbons and<br />

alkylating agents. However, these methods involve a couple <strong>of</strong> time<br />

consuming steps for the isolation <strong>of</strong> an appropriate antibody. The<br />

respective haemoglobin adduct has to be chemically synthesised, an animal<br />

has to be immunised with the modified haemoglobin and, later on,<br />

polyclonal antibodies can be isolated from the blood <strong>of</strong> the immunised<br />

animal. In order to produce monoclonal antibodies, which normally have a<br />

better specificity and sensitivity, spleen cells <strong>of</strong> the immunised animal are<br />

fused with myeloma cells and the antibodies can be isolated from the cell<br />

culture.<br />

The methods for the determination <strong>of</strong> adducts include competitive<br />

radioimmunoassays and solid phase assays (ELISA, USERIA). The protein<br />

is partially hydrolysed, adsorbed on a solid surface and treated with the<br />

primary antibody. An anti-antibody which is directed against the primary<br />

antibody, radiolabelled, or conjugated to a fluorescent dye or an indicator<br />

enzyme, is added and the amount <strong>of</strong> bound label is quantified (Santella et al.,<br />

1986; Lee and Santella, 1988).<br />

DNA adducts<br />

Physical methods<br />

Aromatic amines and nitroarenes<br />


The hydroxylamines produced by enzymatic hydroxylation <strong>of</strong> aromatic<br />

amines or by reduction <strong>of</strong> nitrosoarenes are further conjugated (Osulphatation,<br />

O-acetylation, O-glucuronidation). The conjugates can<br />

decompose to the respective nitrenium ions which add predominantly to<br />

the C8 <strong>of</strong> guanine.<br />

Similarly to protein adducts <strong>of</strong> these compounds, the adducts can be<br />

liberated from DNA by alkaline hydrolysis or hydrazinolysis, extracted and


quantified by HPLC with fluorescence or electrochemical detection. In<br />

order to improve the sensitivity the extracted adducts can be derivatised<br />

with electrophores and analysed by GC/MS (Bakthavachalam et al., 1991;<br />

Lin et al., 1991).<br />

Polycyclic aromatic hydrocarbons<br />

The diol epoxides enzymatically produced from polycyclic aromatic<br />

hydrocarbons mainly adduct at the exocyclic amino group <strong>of</strong> guanine. The<br />

adducts can be liberated from DNA by acid hydrolysis, extracted and<br />

quantified by HPLC with fluorescence or electrochemical detection or by<br />

GC with electron capture detection or GC/MS after suitable derivatisation<br />

with electrophores (Rahn et al., 1982; Shugart and Kao, 1985; Weston et<br />

al., 1989).<br />

Alkylating agents<br />

Alkylating agents mainly alkylate the N7 <strong>of</strong> guanine but also give rise to<br />

the formation <strong>of</strong> other N- and O-alkyl nucleobase adducts. The DNA bases<br />

are liberated by hydrolysis and analysed for the presence <strong>of</strong> adducts by<br />

HPLC with electrochemical detection or they are extracted, derivatised<br />

with electrophores and analysed by GC/MS (Minnetian et al., 1987; Groot<br />

et al., 1994). Some alkylating agents and small epoxides lead to the<br />

formation <strong>of</strong> cyclic nucleobase adducts which exhibit strong fluorescence.<br />

Enzymatic or acid hydrolysis can be used for the liberation <strong>of</strong> the DNA<br />

constituents and the fluorescent adducts can be analysed by HPLC with<br />

fluorescence detection (Fedtke et al., 1990; Steiner et al., 1992a).<br />

Immunological methods<br />

Immunological methods for the determination <strong>of</strong> DNA adducts essentially<br />

follow the procedure as outlined already for protein adducts: generation <strong>of</strong><br />

an antibody, absorption <strong>of</strong> the DNA on a solid surface, incubation with the<br />

antibody and a labelled anti-antibody. However, for the production <strong>of</strong> the<br />

antibody an additional step has to be performed. The immune system<br />

normally does not respond to small molecules. Therefore, the chemically<br />

synthesised base or nucleoside adduct has to be coupled to a carrier protein,<br />

in order to obtain an immunogen (Perera et al., 1986; Santella, 1988;<br />

Poirier, 1993).<br />

Postlabelling<br />

One <strong>of</strong> the most popular assays for determination <strong>of</strong> DNA adducts is the<br />

postlabelling assay. The DNA is enzymatically hydrolysed to the four

natural deoxynucleoside-3′-monophosphates (dNp) and the dNp adducts.<br />

The adducted dNp carrying bulky or aromatic substituents are enriched by<br />

extraction with butanol in the presence <strong>of</strong> a phase transfer agent or by<br />

selective digestion <strong>of</strong> the natural (unadducted) dNp with nuclease P1. The<br />

enriched adducted dNp are labelled with [ 32 P]- or [ 33 P]ATP (Figure 6.4).<br />

Polynucleotide kinase T4 is used to catalyse the transfer <strong>of</strong> the labelled<br />

phosphate group from ATP to the 5′ position <strong>of</strong> the dNp. The labelled<br />

deoxynucleoside-3′,5′-bisphosphates are then separated by multidirectional<br />

TLC on polyethyleneimine coated cellulose. Radioactive impurities and<br />

unused ATP are running to the top with phosphate buffer (D1), whereas<br />

nucleotides carrying aromatic or bulky adducts are retained at or near the<br />

origin. The part containing the impurities and unused ATP is cut <strong>of</strong>f, and<br />

the adducts are chromatographed in D3 (opposite to D1) and D4<br />

(perpendicular to D3) with ammonia and ammonia/ propanol or urea<br />

containing buffers. Adduct spots are visualised and quantified by<br />

autoradiography and Cherenkov counting or by phosphor imaging (Gupta,<br />

1985; Reddy and Randerath, 1986; Beach and Gupta, 1992).<br />

Alternatively, the enriched nucleotide adducts can be chemically<br />

derivatised with fluorescent labels and analysed by HPLC with fluorescence<br />

detection. However, this method does not reach the sensitivity <strong>of</strong> the<br />

radioactive assay (Sharma and Jain, 1991; Jain and Sharma, 1993).<br />

Comparison <strong>of</strong> different methods<br />


The methods used for the determination <strong>of</strong> protein and DNA adducts are<br />

summarised in Tables 6.1 and 6.2. Special attention is drawn to the cost <strong>of</strong><br />

equipment and time required for analysis.<br />

HPLC methods with electrochemical or fluorescent detection are<br />

relatively insensitive and only applicable with compounds which are<br />

strongly fluorescent or electrochemically active. Since the costs for the<br />

equipment used and the time consumption are relatively low, these<br />

methods are attractive in certain cases. GC with electron capture detection<br />

or GC/MS <strong>of</strong>fers better sensitivity. However, the method requires<br />

derivatisation. In addition, the costs for the equipment <strong>of</strong> the GC/MS<br />

methods are quite high. Immunoassays are very sensitive, but involve a<br />

number <strong>of</strong> time consuming steps for the preparation <strong>of</strong> an appropriate<br />

antibody, and are only possible if the structure <strong>of</strong> the respective adduct is<br />

known. The postlabelling method for DNA adducts <strong>of</strong>fers the best<br />

sensitivity, with low equipment costs and low to medium time<br />

consumption. However, the standard method only detects bulky or<br />

aromatic adducts.


Figure 6.4 Schematic representation <strong>of</strong> the postlabelling assay for determination <strong>of</strong><br />

DNA adducts (Gupta, 1985; Reddy and Randerath, 1986; Beach and Gupta, 1992).<br />

Examples/applications<br />

In the following sections some useful applications <strong>of</strong> adduct<br />

determinations, which have been performed in our laboratory, will be<br />


Table 6.1 Methods for the determination <strong>of</strong> protein adducts<br />

a An approximate mean sensitivity is given in pmol (=10 −12 mol) adducts/g<br />

haemoglobin.<br />

Table 6.2 Methods for the determination <strong>of</strong> DNA adducts<br />

Lack <strong>of</strong> bioavailability <strong>of</strong> 3,3′-dichlorobenzidine from<br />

diarylide pigments<br />


a An approximate mean sensitivity is given in fmol (=10 −15 mol) adducts/mg DNA.<br />

3,3′-Dichlorobenzidine is an important intermediate in the production <strong>of</strong><br />

diarylide pigments and azo dyes. Some <strong>of</strong> these pigments have been tested<br />

in long term studies and shown to exert no specific toxicological effects and<br />

to be not carcinogenic to experimental animals (ETAD Report, 1990).<br />

However, there might be a theoretical hazard after metabolic splitting <strong>of</strong> the<br />

pigments into DCB, a known animal carcinogen (IARC, 1982). DCB and<br />

its N-acetylated metabolite are N-hydroxylated and oxidised to the<br />

corresponding nitroso compound which binds to haemoglobin. Since no<br />

repair <strong>of</strong> haemoglobin adducts occurs, these adducts cumulate during the


Figure 6.5 HPLC/ECD pr<strong>of</strong>iles obtained after hydrolysis and extraction <strong>of</strong><br />

haemoglobin samples isolated from an untreated rat (control) and from rats treated<br />

for 4 weeks with DCB (2 mg kg −1 ), Direct Red 46 (160 mg kg −1 ), Pigment Yellow<br />

13 (400 mg kg −1 ) and Pigment Yellow 17 (400 mg kg −l ) as well as <strong>of</strong> commercially<br />

available bovine haemoglobin (Hb-bovine).<br />

life span <strong>of</strong> the erythrocyte. Haemoglobin adduct formation, therefore, was<br />

used to monitor the liberation <strong>of</strong> DCB from diarylide pigments.<br />

Rats were treated by daily oral gavage for 4 weeks with the pigment at<br />

daily dose levels <strong>of</strong> 400 mg kg −1 body weight. As a positive control,<br />

animals were treated accordingly with DCB (2 mg kg −1 ) or with Direct Red<br />

46 (160 mg kg −1 ), asoluble azo dye with known bioavailability <strong>of</strong> DCB.<br />

After termination <strong>of</strong> the treatment, haemoglobin was isolated and<br />

hydrolysed in 0.1 N sodium hydroxide. The liberated DCB and<br />

monoacetyl-DCB were extracted with toluene/2-propanol and analysed by<br />

HPLC with electrochemical detection. With 2 mg DCB kg −1 body weight<br />

DCB and monoacetyl-DCB adducts were clearly detectable, amounting up

to 50 ng g −1 haemoglobin (Figure 6.5). No macromolecular adducts were<br />

detectable in the rats treated with the two diarylide pigments. The limits <strong>of</strong><br />

determination would correspond to a daily DCB dose <strong>of</strong> 0.3–0.5 mg kg −1<br />

body weight, indicating that DCB was not liberated from the pigments at a<br />

determination limit <strong>of</strong> 0.3% <strong>of</strong> the DCB equivalents, whereas the<br />

bioavailability <strong>of</strong> DCB in the rats treated with the azo dye could clearly be<br />

confirmed.<br />

Formation <strong>of</strong> glycidaldehyde from glycidylethers<br />

Bisphenol A diglycidylether (BPADGE) is widely used as component <strong>of</strong><br />

epoxy resins. The chemical reactivity <strong>of</strong> this class <strong>of</strong> compounds is a<br />

prerequisite for their technical use, and accounts for the sensitising,<br />

mutagenic and in some cases carcinogenic properties <strong>of</strong> many epoxy resin<br />

monomers. It was suggested that the metabolic inactivation <strong>of</strong> BPADGE by<br />

hydrolysis <strong>of</strong> epoxides may form an equilibrium with its metabolic<br />

activation by oxidative dealkylation <strong>of</strong> the intact glycidyl side chain<br />

followed by the release <strong>of</strong> glycidaldehyde. Cutaneous treatment <strong>of</strong> mice<br />

with glycidaldehyde led to the formation <strong>of</strong> one major epidermal DNA<br />

adduct which was identified as HMEdA<br />

(hydroxymethylethenodeoxyadenosine, Steiner et al., 1992a). This cyclic<br />

deoxyadenosine adduct is strongly fluorescent and can be quantified by<br />

fluorescence measurements.<br />

In order to investigate the formation <strong>of</strong> glycidaldehyde from BPADGE,<br />

mice were treated with BPADGE (2 mg) and the fluorescent<br />

glycidaldehydeDNA adducts formed in epidermal DNA were compared<br />

with those obtained after treatment with glycidaldehyde (2 mg). After 24–<br />

96 h epidermal DNA was isolated, enzymatically digested to the<br />

deoxynucleoside-3'-monophosphates and analysed for the presence <strong>of</strong><br />

HMEdA by HPLC with fluorescence detection (excitation at 231 nm,<br />

emission at 420 nm). In glycidaldehyde treated mice 166 adducts per 10 6<br />

nucleotides could be detected after an exposure time <strong>of</strong> 24 h (Figure 6.6)<br />

whereas with epidermal DNA from BPADGE treated mice 0.2– 0.8<br />

adducts per 10 6 nucleotides were found. This adduct level would be equal<br />

to a dose <strong>of</strong> 10 µg glycidaldehyde, indicating that, at the most, 1.1% <strong>of</strong> the<br />

glycidaldehyde moiety in BPADGE were bioavailable for DNA-adduct<br />

formation (Steiner et al., 1992b).<br />

Determination <strong>of</strong> reactive compounds in unknown<br />

mixtures<br />


A challenging task is the analysis <strong>of</strong> reactive metabolities in unknown<br />

mixtures <strong>of</strong> different compounds. In order to assess the impact <strong>of</strong> chemical<br />

pollution on aquatic organisms, rainbow trouts were continuously exposed


to the diluted effluent discharges <strong>of</strong> a chemical production plant for 3<br />

months. The plant produced different dyes and chemicals and the waste<br />

water therefore could be contaminated with a variety <strong>of</strong> aliphatic and<br />

aromatic amines and some cyclic aromatic hydrocarbons. After termination<br />

<strong>of</strong> the treatment, liver and gill DNA from exposed and control trouts was<br />

analysed by [ 32 P]postlabelling for the presence <strong>of</strong> DNA adducts.<br />

The DNA was enzymatically hydrolysed to the nucleotides. Adducted<br />

nucleotides were extracted with butanol in the presence <strong>of</strong> the phase<br />

transfer agent tetrabutylammonium chloride and postradiolabelled with<br />

[ 32 P]ATP and PNK. The labelled nucleotides were separated by<br />

multidirectional TLC with 1.0 M phosphate buffer, pH 6.6, in D1, 0.4 M<br />

ammonia in D3 and 4 N ammonia/propanol (1.2:1) in D4. A final<br />

development in direction D4 with 1.0 M phosphate buffer, pH 6.6, was<br />

used as background clean up.<br />

In the trouts exposed to control water no DNA adducts were detectable,<br />

neither in the livers nor in the gills (Figure 6.7). In contrast, in the trouts<br />

exposed to the highest concentration <strong>of</strong> the waste water, at least 4 DNA<br />

adducts could be found in the livers and in the gills. The overall DNA<br />

adduct level in the exposed trouts was relatively low (1 adduct per 10 8<br />

nucleotides, which indicated only a minimal cancer risk for the exposed<br />

fish.<br />

Limitations<br />

However, the methods presented for adduct determination have their<br />

limitations. For protein adduct determination the most popular method is<br />

by HPLC with electrochemical or fluorescence detection after hydrolysis<br />

and extraction <strong>of</strong> the adducts. This is due to the low cost and time<br />

consumption <strong>of</strong> the method. This method is hampered by the possibility <strong>of</strong><br />

interferences, which can elute in the range <strong>of</strong> the compounds <strong>of</strong> interest. For<br />

an exclusion <strong>of</strong> haemoglobin adducts formation at low levels it is therefore<br />

crucial to obtain additional information about the chromatographic peaks<br />

<strong>of</strong> interest, such as for example, by GC/MS.<br />

DNA adducts are <strong>of</strong>ten assessed by [ 32 P]postlabelling. This method is<br />

limited by low yields <strong>of</strong> the enrichment and labelling procedures and by<br />

choosing the appropriate chromatographic conditions for the resolution <strong>of</strong><br />

the labelled adducts. The lack <strong>of</strong> detectability <strong>of</strong> some DNA adducts,<br />

although they may contain aromatic moieties, enforces the use <strong>of</strong> a positive<br />

standard in order to check for the yield <strong>of</strong> the enrichment and the labelling<br />

reaction, and to check for appropriate chromatographic conditions to<br />

resolve the adducts.


Figure 6.6 HPLC/fluorescence analysis <strong>of</strong> epidermal DNA hydrolysates from a<br />

control (a) and a BPADGE treated mouse (b), and UV trace <strong>of</strong> synthetic HMEdAp<br />

and HMEdGp (c).


Figure 6.7 TLC chromatograms <strong>of</strong> DNA adducts in gills and livers <strong>of</strong> rainbow<br />

trouts, exposed for 3 months to waste water or control water. Top: control water,<br />

liver DNA (left chromatogram), gill DNA (right chromatogram); bottom: waste<br />

water, liver DNA (left chromatogram), gill DNA (right chromatogram).<br />

Conclusions<br />

Each method, although inherently chemically-specific, has its advantages<br />

and limitations depending on the adduct-type. The continued rapid<br />

development <strong>of</strong> the technologies described for assessing biomarkers should<br />

result in more accurate assessment <strong>of</strong> the intracellular reactions <strong>of</strong><br />

chemicals and thereby provide information about the mechanism <strong>of</strong><br />

toxicity <strong>of</strong> a compound under investigation.

Acknowledgements<br />

Grateful thanks to Drs Markus Joppich and Regula Joppich-Kuhn for<br />

haemoglobin adduct analyses and Dr Sandra Steiner for the development<br />

<strong>of</strong> the fluorescence assay for HMEdAp.<br />

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HARRIS, C.C., 1989, Fluorescence and mass spectral evidence for the<br />

formation <strong>of</strong> benzo[a]pyrene anti-diol-epoxide-DNA and -haemoglobin<br />

adducts in humans, Carcinogenesis, 10, 251–7.


Pulmonary toxicology <strong>of</strong> industrial<br />


7<br />

Studies to Assess the Carcinogenic Potential <strong>of</strong><br />

Man-Made Vitreous Fibers<br />



Schuller International, Inc., Littleton, CO<br />

Introduction<br />

Man-made vitreous fibers (MMVFs) are a class <strong>of</strong> materials which have<br />

found many applications in both residential and industrial settings. MMVFs<br />

are fibrous inorganic substances that are made primarily from rock, clay,<br />

slag or glass. Sometimes referred to as man-made mineral fibers<br />

(MMMFs), the major classes <strong>of</strong> MMVF are refractory ceramic fibers<br />

(RCFs), fibrous glass, rock (stone) wool and slag wool.<br />

RCF, the smallest category <strong>of</strong> MMVF, represents only about 1–2 per<br />

cent <strong>of</strong> the world production <strong>of</strong> MMVF. It is made by melting Al 2O 3 and<br />

SiO 2 in about equal amounts or by melting kaolin clay and then ‘spinning’<br />

or ‘blowing’ this molten material into fibers. Most RCF is used as a high<br />

temperature furnace insulation. World production <strong>of</strong> RCF in 1990 was<br />

about 80 million 1b. Fibrous glass is the largest category <strong>of</strong> the MMVFs<br />

and is used in insulation, air handling, filtration and sound absorption. The<br />

thermal, acoustical and fire resistant properties <strong>of</strong> these products have led<br />

to their widespread use in a variety <strong>of</strong> residential and commercial<br />

applications. Production <strong>of</strong> fibrous glass in North America in 1989 was<br />

approximately 1.8 million t. Slag and rock wool are composed primarily <strong>of</strong><br />

calcium, magnesium, aluminum and silica. Since 1975, most slag wool has<br />

been produced from the waste slag that resulted from the reduction <strong>of</strong> iron<br />

ore to iron. Rock wool fibers are made from basaltic rocks with additives<br />

such as limestone or dolomite. Slag and rock wool are used in residential<br />

and commercial low and high temperature insulation and in acoustical<br />

ceiling tiles and wall panels. About 75% <strong>of</strong> slag wool production is used in<br />

acoustical ceiling tile manufacture in North America.<br />

Animal studies and epidemiological studies have been conducted to<br />

assess the potential biological effects <strong>of</strong> MMVFs. This research has been<br />

reviewed by the International Agency for Research on Cancer (IARC, 1988),<br />

the International Programme on Chemical Safety (IPCS, 1988), and the US<br />

Environmental Protection Agency (Vu, 1988). These reviews are consistent


in the judgement that chronic inhalation studies <strong>of</strong> airborne fibers provide<br />

the best model for assessing potential risk to man (McClellan et al., 1992).<br />

In assessing the carcinogenic risks <strong>of</strong> exposure to any possible<br />

occupational hazard, research is pursued through several different scientific<br />

techniques. Studies <strong>of</strong> mortality (analysis <strong>of</strong> death rates) are used to<br />

evaluate the potential carcinogenicity associated with direct human<br />

exposure. Animal exposure studies are used to not only evaluate the<br />

potential carcinogenicity but to also investigate the mechanisms <strong>of</strong> disease<br />

development. <strong>Industrial</strong> hygiene and engineering studies are used for<br />

quantifying exposures.<br />

Epidemiological studies<br />

By general agreement among experts (IARC, 1988; IPCS, 1988), two major<br />

historical cohort studies are considered to have comprehensively addressed<br />

the mortality experience <strong>of</strong> workers engaged in the production <strong>of</strong> FG, rock<br />

wool and slag wool: a European study conducted by the International<br />

Agency for Research on Cancer (IARC), and a University <strong>of</strong> Pittsburgh<br />

study conducted in the USA. The discussion here will concentrate on those<br />

two studies. For a summary <strong>of</strong> other studies, the reader is referred to the<br />

IARC review (IARC, 1988). There are no published reports <strong>of</strong> the<br />

mortality experience <strong>of</strong> RCF workers. Epidemiological studies <strong>of</strong> workers<br />

engaged in the manufacture <strong>of</strong> all major classes <strong>of</strong> MMVF are underway or<br />

are continuing. Morbidity studies <strong>of</strong> the respiratory health <strong>of</strong> workers are<br />

not discussed here.<br />

The IARC study<br />

IARC researchers reported their study at the WHO Occupational Health<br />

Conference on the Biological Effects <strong>of</strong> Man-Made Mineral Fibres at<br />

Copenhagen in 1982, with a follow-up in 1986 (Simonato et al., 1987). The<br />

updated study is also published in the Scandinavian Journal <strong>of</strong> Work,<br />

Environment & Health, Volume 12, Supplement 1, 1986. The mortality <strong>of</strong><br />

23609 workers (2836 deaths) employed in 13 European factories engaged<br />

in the production <strong>of</strong> MMVF (including 11 852 fibre glass production<br />

workers at six plants in five countries and 10 115 rock wool/slag wool<br />

production workers at seven plants in four countries) has been studied<br />

(Saracci et al., 1984) and updated (Simonato et al., 1987). The authors<br />

reported an ‘excess <strong>of</strong> lung cancer among rock-wool/slag workers<br />

employed during an early technological phase before the introduction <strong>of</strong><br />

dust-suppressing agents’, and concluded that ‘fiber exposure, either alone or<br />

in combination with other exposures, may have contributed to the elevated<br />

risk’. The authors also reported that ‘no excess <strong>of</strong> the same magnitude was<br />

evident for glass-wool production, and the follow-up <strong>of</strong> the continuous-

filament cohort was too short to allow for an evaluation <strong>of</strong> possible longterm<br />

effects’. It was also noted that ‘there was no evidence <strong>of</strong> an increased<br />

risk for pleural tumors or non-malignant respiratory diseases’. An update <strong>of</strong><br />

this study is underway.<br />

The University <strong>of</strong> Pittsburgh study<br />

This study was also reported at the WHO Occupational Conference on<br />

Biological Effects <strong>of</strong> Man-Made Mineral Fibres at Copenhagen in 1982 and<br />

the follow-up Conference in 1986 (Simonato et al., 1987). Subsequent to<br />

the 1986 Conference, additional analyses were completed and included in<br />

the manuscript published for the proceedings (Enterline et al., 1987). The<br />

study has been updated and published (Marsh et al., 1990). The University<br />

<strong>of</strong> Pittsburgh researchers’ comprehensive mortality review <strong>of</strong> more than<br />

16000 workers— many with long-term exposure up to 40 years—was<br />

undertaken at 17 US fiber-glass, rock-wool and slag-wool manufacturing<br />

plants, including 14800 fiber-glass workers in 11 plants. The original<br />

report, given in 1982, covered the mortality experience from the 1940s to<br />

the end <strong>of</strong> 1977. The same group <strong>of</strong> workers was followed through 1982<br />

(reported in October 1986, with additional analyses available in June<br />

1987). The June 1987 report contained, for the first time, local area<br />

mortality statistics for each <strong>of</strong> the plants as the basis for studying the<br />

mortality experience. Experts agree that, barring unusual circumstances,<br />

local area comparisons are most appropriate. The study has been further<br />

updated through 1985, with publication in 1990. For respiratory cancer, in<br />

the latest update there was a small but statistically significant increase for<br />

fiber glass production workers. However, aside from the issue <strong>of</strong><br />

uncontrolled potential confounding, the study provides no evidence to date<br />

that respiratory cancer mortality is related to fiber glass exposure. There<br />

was a somewhat larger statistically significant excess <strong>of</strong> respiratory cancer<br />

mortality reported for slag wool and rock wool production workers. The<br />

absence <strong>of</strong> any clear exposure-response relationship for any <strong>of</strong> the fiber<br />

groups studied led the authors to conclude that ‘overall, the evidence <strong>of</strong> a<br />

relationship between exposure to man-made mineral fibers and respiratory<br />

cancer appears to be somewhat weaker than in the previous update’.<br />

Consistent with the IARC study, no increase in the occurrence <strong>of</strong><br />

mesothelioma has been observed in this cohort. This study has now been<br />

expanded to include well over 30000 workers from 14 fiber-glass and six<br />

rock wool and slag wool facilities.<br />

Other epidemiological studies<br />


In addition to the two major studies highlighted above, a number <strong>of</strong> other<br />

studies have been conducted as well. Many <strong>of</strong> them widely overlap these


major studies, comprise sub-groups within them, or represent smaller<br />

worker populations outside <strong>of</strong> them.<br />

A Canadian study was reported by Shannon at the WHO Occupational<br />

Health Conference on Biological Effects <strong>of</strong> Man-Made Mineral Fibers at<br />

Copenhagen in 1982 and 1986 (Shannon et al., 1987). It followed 2557<br />

male workers at a Canadian glass wool plant through 1977 and was later<br />

updated to extend the follow-up to the end <strong>of</strong> 1984. In the updated study,<br />

the authors reported a statistically significant excess <strong>of</strong> lung cancer. In<br />

discussing this excess, the authors concluded that the interpretation <strong>of</strong> the<br />

information was difficult since there was no relationship between the<br />

excess <strong>of</strong> lung cancer and the length <strong>of</strong> time since first exposure to the<br />

fibrous glass manufacturing environment.<br />

Two recent case-control studies have addressed the lung cancer mortality<br />

<strong>of</strong> FG and slag wool production workers. Chiazze et al. (1992) have<br />

investigated the potential impact <strong>of</strong> confounding factors such as smoking<br />

and other occupational exposures for workers at the oldest and largest US<br />

fiber glass manufacturing facility. In particular, Chiazze helped clarify the<br />

heavy smoking patterns in those workers and verified the large impact that<br />

smoking has on their lung cancer experience. Wong et al., (1991)<br />

investigated the potential impact <strong>of</strong> smoking on the lung cancer deaths at<br />

nine US slag wool manufacturing plants. Wong also found heavy smoking<br />

among the slag wool workers and advanced the understanding <strong>of</strong> the<br />

modest increase in lung cancer seen in the historical cohort studies cited<br />

above.<br />

Users <strong>of</strong> MMVFs generally have experienced mixed exposures, making<br />

the study <strong>of</strong> any potential health effects <strong>of</strong> MMVF difficult, if possible at<br />

all. For example, in a study <strong>of</strong> Swedish construction workers, Engholm et al.<br />

(1987) discussed the difficulty caused by overlapping <strong>of</strong> reported exposures<br />

to asbestos and MMVFs. In addition, essential employment and exposure<br />

histories for users <strong>of</strong> MMVFs are lacking.<br />

The mortality studies <strong>of</strong> FG workers, while showing a small but<br />

statistically significant increase in lung cancer, have failed to show any<br />

consistent relationship with exposure to FG (i.e. no dose-response<br />

relationships have been found). It is recognized that uncontrolled<br />

occupational and/or non-occupational confounding factors may be<br />

associated with the slight increase. The IARC review (IARC, 1988)<br />

concluded that there is ‘inadequate evidence’ for carcinogenicity in<br />

humans. Other reviews have reached similar conclusions. In addition,<br />

reports subsequent to the IARC review have further clarified potential<br />

confounding factors and, if anything, shown weaker evidence <strong>of</strong> a<br />

relationship between exposure and lung cancer.<br />

The cohort mortality studies <strong>of</strong> rock wool and slag wool workers have<br />

shown a somewhat larger statistically significant excess <strong>of</strong> lung cancer<br />

deaths, but have also provided no clear dose-response relationship with

fiber exposure. While the IARC review (IARC, 1988) concluded that there<br />

is ‘limited evidence’ for carcinogenicity in humans, reports subsequent to<br />

the IARC review have further clarified potential confounding factors and,<br />

if anything, shown weaker evidence <strong>of</strong> a relationship between exposure and<br />

lung cancer.<br />

Experimental studies<br />

Toxicologic studies <strong>of</strong> MMVFs have been conducted in both in vitro and in<br />

vivo systems. In addition, the physical and chemical characteristics thought<br />

to correlate with toxicity have been examined. The in vitro studies have<br />

been conducted using cells from the lungs <strong>of</strong> animals as well as bacterial<br />

and cell lines. Two categories <strong>of</strong> whole-animal studies have been reported:<br />

studies using artificial methods to implant high concentrations <strong>of</strong> fibers in<br />

the abdomen, pleura or trachea <strong>of</strong> animals; and inhalation studies <strong>of</strong><br />

maximum tolerated doses and multiple dose levels <strong>of</strong> fibers.<br />

Cell culture studies<br />


The use <strong>of</strong> cell culture systems for studying the toxic effects <strong>of</strong> fibers has<br />

been recently reviewed (Hesterberg et al., 1993a). A number <strong>of</strong> studies<br />

have shown that fiber length and diameter are important in determining<br />

the toxicity <strong>of</strong> mineral fibers <strong>of</strong> various chemical compositions to cells<br />

grown in culture (Chamberlain et al., 1979, Tilkes and Beck, 1980;<br />

Hesterberg and Barrett, 1984; Hesterberg et al., 1993a; Hart et al., 1994).<br />

Chemical composition has also been shown to be critical to the toxicity <strong>of</strong><br />

fibers to rat tracheal epithelial cells (Ririe et al., 1985) and human<br />

bronchial epithelial cells grown in culture (Kodama et al., 1993). MMVFs<br />

have also been shown to induce neoplastic transformation (Hesterberg and<br />

Barrett, 1984; Poole et al., 1986) and genetic damage to cells in culture<br />

(Sincock and Seabright, 1975; Oshimura et al., 1984). Cell culture models<br />

are important for understanding the mechanisms <strong>of</strong> fiber toxicity and, with<br />

further development, have potential for use as part <strong>of</strong> a battery <strong>of</strong> shortterm<br />

screening tests to assess the toxic and tumorigenic potential <strong>of</strong><br />

mineral fibers. However, it was recently shown that cytotoxicity <strong>of</strong> different<br />

compositions MMVFs to Chinese hamster ovary (CHO) cells in culture did<br />

not correlate with the in vivo toxicity <strong>of</strong> theses MMVFs (Hart et al., 1994).<br />

This may be related to CHO cells being aneuploid, preneoplastic and not a<br />

normal target cell for fiber toxicity in vivo. Future in vitro studies <strong>of</strong> MMVF<br />

toxicity should focus on the use <strong>of</strong> cell types that represent the relevant<br />

target tissues, and cells should be as close to normal as possible.


Implantation studies<br />

Using various types and dimensions <strong>of</strong> fibers, researchers have studied the<br />

effects <strong>of</strong> ‘artificially’ exposed animals by surgically implanting fibrous<br />

material in the pleural (chest) and abdominal cavities <strong>of</strong> laboratory<br />

animals, and by injecting fibers directly into the trachea (Pott et al., 1987;<br />

Stanton et al., 1981). Those studies have shown that high levels <strong>of</strong> most<br />

fibrous materials <strong>of</strong> certain dimensions, regardless <strong>of</strong> their physical or<br />

chemical makeup, can induce tumors in laboratory animals. From these<br />

study results, scientists have also hypothesized that biological activity<br />

correlates with fiber length and diameter, since ‘long, thin’ fibers are the<br />

most active. The actual chemical composition appears to play only a minor<br />

role, if any, in such ‘artificial exposure’ experiments (Stanton et al., 1981).<br />

Injection <strong>of</strong> fibers bypasses the normal defense mechanisms <strong>of</strong> the lung<br />

and can produce abnormal fiber distribution, fiber clumping, and overload<br />

doses (McClellan et al., 1992). Furthermore, when fibers are injected into<br />

the pleura or peritoneum <strong>of</strong> an animal, leaching, degradation,<br />

fragmentation or any other transformations are unlikely to be the same as<br />

after inhalation. The weaknesses <strong>of</strong> intracavitary injection studies <strong>of</strong><br />

fibrous materials limit their relevance for human risk assessment (IPCS,<br />

1988; Vu, 1988; Dement et al., 1990; WHO, 1992; McClellan et al.,<br />

1992).<br />

Recent animal inhalation studies <strong>of</strong> MMVFs<br />

Inhalation is the only natural route <strong>of</strong> exposure for fiber entry and<br />

distribution to the target organs in man. Animal inhalation studies are<br />

more relevant than intracavity administration studies for risk assessment<br />

because the exposure conditions <strong>of</strong> inhalation experiments more closely<br />

approach the circumstances <strong>of</strong> human exposure.<br />

Background<br />

In June 1988, a series <strong>of</strong> inhalation studies was initiated at Research and<br />

Consulting Company (RCC) in Geneva, Switzerland, to evaluate the<br />

biological effects <strong>of</strong> different compositions <strong>of</strong> MMVF. These included<br />

RCFs, common insulation fiber glass, and rock and slag wool fibers. These<br />

studies used recently perfected state-<strong>of</strong>-the-art technologies for fiber sizeseparation,<br />

fiber l<strong>of</strong>ting and nose-only inhalation exposure. A more<br />

detailed description <strong>of</strong> the techniques used and the results from these<br />

studies are found elsewhere (Hesterberg et al., 1991, 1993b; Mast et al.,<br />

1995 McConnell et al., 1994). The animal models selected were those with<br />

demonstrated capacity to develop asbestosrelated disease following<br />

inhalation exposure. The studies were conducted in accordance with

standard techniques for chronic toxicity/carcinogenicity studies, including<br />

dose and latency considerations. Rats were exposed for 2 years and<br />

hamsters for 18 months. The animals were observed for their lifetime or until<br />

20% survival <strong>of</strong> the test group was reached. Positive and shamexposed<br />

negative controls were included in the protocol.<br />

Fiber aerosols<br />

In designing these animal inhalation studies, the techniques <strong>of</strong> fiber<br />

preparation, aerosolization, exposures, measurement, quantification and<br />

determination <strong>of</strong> actual target organ dose were critical factors. Fiber<br />

dimensions that permitted deposition into the distal lung regions (i.e.<br />

respirable fibers) for the model used were selected. The characteristics <strong>of</strong> the<br />

fiber aerosol in actual work areas for man was an important consideration<br />

in determining experimental exposure. For example, an average fiber size<br />

<strong>of</strong> 1×20 µm has been measured during simulated RCF work practices. The<br />

critical need to use fibers pre-selected for their size and to verify the actual<br />

size distributions <strong>of</strong> the fiber exposure aerosol was met throughout the<br />

study. Non-fibrous particles (shot) in the aerosol were reduced to the<br />

maximum extent possible. Furthermore, fiber preparation, handling and<br />

aerosolization did not alter the physical-chemical characteristics <strong>of</strong> the<br />

fiber, since as will be discussed later, these are known to be critical<br />

determinants <strong>of</strong> fiber toxicity.<br />

Nose-only rather than whole-body exposure was used for several<br />

reasons, including the impossibility <strong>of</strong> preparing the huge quantities <strong>of</strong><br />

specially sized fibers that would be required for 2 years <strong>of</strong> whole-body<br />

exposure. Additionally, nose-only exposure levels permitted better control<br />

<strong>of</strong> exposure levels and host entry.<br />

Selection <strong>of</strong> exposure concentrations<br />

It was important that at least three exposure concentrations be used in the<br />

chronic inhalation study in order to assess the dose-response relationships<br />

<strong>of</strong> any induced changes. The highest concentration selected was the<br />

‘Maximum tolerated dose’ (MTD), while lower concentrations were 50 per<br />

cent <strong>of</strong> the MTD and multiples <strong>of</strong> the projected occupational and<br />

environmental exposure levels.<br />

Experimental design, time lines<br />


Groups <strong>of</strong> three or six randomly selected animals from each exposure<br />

group were killed at 3, 6, 12, 18 and 24 (rats only) months to follow the<br />

progression <strong>of</strong> histopathological changes and to determine lung fiber<br />

burdens. An additional six ‘recovery’ animals were removed from each


exposure group at 3, 6, 12 and 18 (rats only) months and held without<br />

further treatment until the end <strong>of</strong> the exposure period, when they were<br />

killed to assess progression or regression <strong>of</strong> lung lesions and lung retention<br />

and clearance <strong>of</strong> fibers after cessation <strong>of</strong> exposure. To assure quality<br />

control, the l<strong>of</strong>ting technique and exposure level were consistently<br />

monitored during the study by both gravimetric measurement and fiber<br />

counting techniques. The terminal sacrifice was carried out when only 20 per<br />

cent <strong>of</strong> the animals survived. A complete necropsy was performed on each<br />

animal. Gross pathological examination and diagnoses were performed<br />

using a dissecting microscope. Uniform sections <strong>of</strong> the left lung and right<br />

diaphragmatic lobe were embedded in paraffin, cut at a thickness <strong>of</strong> 4 mm<br />

and replicate sections were routinely stained with hematoxylin and eosin<br />

(H&E) and Masson-Goldner’s trichrome stain for collagen staining to<br />

assess the presence <strong>of</strong> lung fibrosis. In addition, sections were made from<br />

all grossly visible lesions from that and other portions <strong>of</strong> the lung.<br />

Proliferative lesions <strong>of</strong> the pulmonary parenchyma were designated as<br />

bronchoalveolar hyperplasia, pulmonary adenoma or adenocarcinoma.<br />

Other types <strong>of</strong> lesions, including those in the pleura were noted where<br />

appropriate. All research and analyses were conducted using good<br />

laboratory practices.<br />

Lung fiber burden<br />

Immediately after necropsy, the infracardiac lobe <strong>of</strong> each animal’s lung was<br />

removed and frozen for later analysis <strong>of</strong> lung fiber burden. To recover<br />

fibers from the lung, the tissue was rapidly dehydrated with acetone and<br />

ashed using a low-temperature process. Recovered fibers were dispersed in<br />

distilled water and examined using scanning electron microscopy. Number,<br />

dimensions and other physical characteristics <strong>of</strong> the inhaled lung fibers<br />

were determined, and reported as fibers per mg <strong>of</strong> dry lung weight.<br />

Results from recent animal inhalation studies <strong>of</strong> MMVFs<br />

Refractory ceramic fibers<br />

In the first RCC studies, rodents were exposed to the MTD <strong>of</strong> the sizeselected<br />

RCF test fiber, 30 mg m −3 and approximately 200–250 fibers cm<br />

−3 . Rats were exposed for 6 h per day, 5 days a week to aerosols containing<br />

one <strong>of</strong> four different types <strong>of</strong> RCF: kaolin, RCF 1; zirconia, RCF 2; high<br />

purity kaolin, RCF 3; and ‘after service’ (a kaolin based ceramic fiber<br />

containing 27% crystalline silica that had previously been exposed to high<br />

temperature), RCF 4. Hamsters were exposed to only kaolin RCF fibers.<br />

Positive controls (chrysotile asbestos) and negative controls (filtered air)

Table 7.1 Summary <strong>of</strong> lung pathology findings in RCF hamster inhalation study<br />

a WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter


Table 7.2 Summary <strong>of</strong> lung pathology findings in RCF MTD (30 mg m−3) rat<br />

inhalation<br />

a WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter

asbestos (10 mg m −3 ) were included in the study. The crocidolite exposure<br />

had to be stopped at 10 months due to excessive mortality resulting from<br />

lung toxicity. The results are summarized in Table 7.5. Crocidolite<br />

exposure resulted in lung fibrosis, a significant increase in lung tumors, and<br />

a single mesothelioma. Rock wool, but not slag wool, exposure at 16 and<br />

30 mg m −3 resulted in minimal lung fibrosis. However, neither rock wool<br />

nor slag wool exposure resulted in mesotheliomas or a significant increase<br />

in lung tumors.<br />

Lung burden analyses<br />


Table 7.3 Combined summary <strong>of</strong> lung pathology findings: chrysotile and RCF1<br />

from RCF MTD study in rats; and RCF multidose study in rats<br />

a WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter


Table 7.4 Summary <strong>of</strong> lung pathology findings in fibrous glass inhalation study in<br />

rats<br />

a WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter 5 µm, and a diameter 10 µm in length in the lung were similar for each <strong>of</strong> the<br />

different MMVF types (Figures 7.2(a) and 7.2(b)). However, greater<br />

numbers <strong>of</strong> long fibers (>20 µm. long) were found in the lungs <strong>of</strong> rats<br />

exposed to RCF 1 and MMVF 21 (rock wool) than for the other fiber<br />

types (Figure 7.2(c)). Even though lung levels <strong>of</strong> long MMVF 21 fibers<br />

were higher than long RCF 1 fibers, lung fibrosis occurred much later for<br />

MMVF 21 (18 vs 6 months for RCF 1) and no mesotheliomas or significant<br />

increase in lung tumors were observed for MMVF 21. This indicates that<br />

the lung pathogenic potential <strong>of</strong> a fiber may be determined by more than<br />

dose and dimension.


Table 7.5 Summary <strong>of</strong> lung pathology findings in rock and slag wool inhalation<br />

study in rats<br />

a WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter


disappearance <strong>of</strong> long fibers. More studies are required to determine if a<br />

fiber’s ability to be leached is a critical determinant <strong>of</strong> its ultimate toxicity<br />

to the lung.<br />

Figure 7.1 Length distributions (a) and diameter distributions (b) <strong>of</strong> fibers from the<br />

lungs <strong>of</strong> rats exposed for 13 weeks to the five different MMVFs in the chronic<br />

inhalation studies. To permit clearance <strong>of</strong> the upper airways, rats were killed a


minimum <strong>of</strong> 24 h after the exposure was stopped; the right accessory lobe was<br />

frozen and later low temperature ashed for fiber recovery. Fiber lengths were<br />

determined using phase contrast optical microscopy, while fiber diameters were<br />

determined using scanning electron microscopy.<br />

Results from previous MMVF inhalation studies<br />

The results from two previous RCF inhalation studies (Davis et al., 1984;<br />

Smith et al., 1987) differ from the more recent RCC studies presented here.<br />

Davis et al., (1984) reported RCF exposure <strong>of</strong> rats resulted in an average <strong>of</strong><br />

5 per cent pulmonary fibrosis, pulmonary tumors in eight <strong>of</strong> 48 rats, and<br />

one peritoneal mesothelioma. The lower fibrosis and tumor response in the<br />

Davis study may have resulted from the lower exposure concentration used;<br />

8.4 mg m −3 compared to 30 mg m −3 in the present study. In addition, the<br />

use <strong>of</strong> fibers that were not presized, the use <strong>of</strong> whole-body exposure, or the<br />

fiber generation technique, which may have crushed some <strong>of</strong> the fibers,<br />

may account for the lack <strong>of</strong> consistency with the present study. Smith et<br />

al., (1987) exposed hamsters and rats to RCF at 200 f cm −3 , 6h a day, 5<br />

days a week, for 24 months. The rat study showed no significant increase<br />

in neoplasms and minimal pulmonary fibrosis in 22% <strong>of</strong> the exposed<br />

animals. In the hamster study, RCF produced only one mesothelioma in 50<br />

animals and no fibrosis was observed. It is difficult to explain why there<br />

was little response to RCF in the Smith studies, but it may be related to the<br />

different aerosol and exposure technology used or to the low exposure<br />

level; 12 mg m −3 compared to 30 mg m −3 in the present study.<br />

Previous inhalation studies <strong>of</strong> FG using rodents agree with the findings<br />

<strong>of</strong> the RCC studies. Fiber glass has been tested by inhalation in guinea pigs<br />

(Gross et al., 1970), hamsters (Lee et al., 1981; Smith et al., 1987), and<br />

rats (Gross et al., 1970; Lee et al., 1981; McConnell et al., 1984; Wagner<br />

et al., 1984; Mitchell et al. 1986; LeBouffant et al., 1987; Muhle et al.,<br />

1987; Smith et al., 1987). None <strong>of</strong> these studies identified a significant<br />

increase in either fibrosis or neoplasms following glass fiber inhalation in<br />

spite <strong>of</strong> FG lung burdens in excess <strong>of</strong> several hundred thousand fibers per<br />

mg dry lung tissue. In three <strong>of</strong> the above studies, the chronic inhalation<br />

toxicity <strong>of</strong> rock and slag wool were also examined (Wagner et al., 1984;<br />

LeBouffant et al., 1987; Smith et al., 1987). As was seen with fibrous glass,<br />

all three studies demonstrated no tumorigenic response by this route <strong>of</strong><br />



Figure 7.2 (continued over) Lung burdens <strong>of</strong> (a) WHO fibers, (b) fibers >10 µm in<br />

length, and (c) fibers >20 µm in length per mg dry lung tissue from the lungs <strong>of</strong> rats<br />

continuously exposed to 30 mg m −3 <strong>of</strong> the five different MMVFs. Rats were killed<br />

at least 24 h after the exposure was stopped, the right accessory lobe was frozen<br />

and later low temperature ashed for fiber recovery.<br />

<strong>Industrial</strong> hygiene studies<br />

RCF<br />

<strong>Industrial</strong> hygiene monitoring data obtained on a regular basis at locations

Figure 7.2 Continued<br />

where RCF products are manufactured show that exposures are generally<br />

below 1.0 f cm −3 , typically below 0.2 f cm −3 . In a recent study, RCF levels<br />

during various end-user operations ranged from 0.12 to 1.55 f cm −3 with<br />

an overall mean and SD <strong>of</strong> 0.74±0.49 f cm −3 (Lees et al., 1993b). Other<br />

end-user studies have indicated that RCF exposures can exceed 5 f cm −3 or<br />

higher if appropriate engineering controls and work practices are not<br />

followed (Schuller, 1985–1988).<br />

Fiber glass<br />


Recently, studies which examined human aerosol exposure to fiber glass in<br />

manufacturing, installation and removal, and in ambient air were reviewed<br />

(Hesterberg and Hart, 1994). In most cases, human exposures to airborne<br />

fiber glass during manufacturing and installation fell well below the OSHAproposed<br />

permissible exposure limit (PEL) <strong>of</strong> 1 f cm −3 air (OSHA, 1992).<br />

Airborne fiber concentrations during FG manufacturing operations are<br />

typically less than 0.2 f cm −3 , with the majority being less than 0.1 f cm −3 .<br />

Exceptions include manufacture <strong>of</strong> finer diameter fiber glass and blowing<br />

installation <strong>of</strong> loose fiber glass that is either milled or lacks binder. Airborne<br />

levels averaging greater than 1 f cm −3 have been reported in the production<br />

<strong>of</strong> finer diameter fiber glass (TIMA, 1990), while blowing installation <strong>of</strong><br />

loose fiber glass without binder resulted in a task length average (TLA) <strong>of</strong><br />

7.67 f cm −3 , and an 8-h TWA <strong>of</strong> 1.96 f cm −3 (Lees et al., 1993a). Blowing<br />

installation <strong>of</strong> loose mineral wool also resulted in higher aerosol levels; a


TLA <strong>of</strong> 1.94 f cm −3 and an 8-h TWA <strong>of</strong> 0.97 f cm −3 (Lees et al., 1993a).<br />

Removal <strong>of</strong> fiber glass insulation created an aerosol <strong>of</strong> 0.042 f cm −3 (Jacob<br />

et al., 1993). Fiber concentrations <strong>of</strong> 0.004 f cm −3 were reported for<br />

buildings recently insulated with FG (Jacob et al., 1992). However this figure<br />

includes all types <strong>of</strong> fibers as it was obtained using optical microscopy. The<br />

background level prior to fiber glass installation was 0.001 f cm −3 .<br />

Ambient environmental exposures to airborne vitreous fibers were<br />

extremely low; exposure levels <strong>of</strong> product-related vitreous fibers reported<br />

for outdoor air was 0.0007 f cm −3 (Tiesler and Draeger, 1994).<br />

In addition to manufacturing and field use surveys, release <strong>of</strong> fibrous<br />

glass during actual use <strong>of</strong> products, particularly fiber released from air<br />

filter media, has been monitored. To determine possible exposure <strong>of</strong><br />

building occupants to fibrous glass, ambient air was sampled in a number<br />

<strong>of</strong> public buildings in which fibrous glass air filtration products had been<br />

installed. These evaluations showed no significant release <strong>of</strong> fibers from the<br />

filters (Balzer et al., 1971; Cholak and Schafer, 1971).<br />

To evaluate the efficiency <strong>of</strong> fibrous glass filter blankets, several high<br />

volume air samples were collected at various points in the ductwork <strong>of</strong> a<br />

large <strong>of</strong>fice complex at the intake and the exhaust prior to changing the<br />

filter media, and at the exhaust 23 days after installation <strong>of</strong> the new filter.<br />

Analyses <strong>of</strong> the samples using electron microscopy indicate little initial<br />

fiber release which decreases rapidly thereafter to the limit <strong>of</strong> detection<br />

(Schuller, 1987).<br />

Rock and slag wool<br />

Airborne concentrations <strong>of</strong> dust and fibers reported from US mineral wool<br />

plants is generally higher than in US glass wool facilities. This includes both<br />

airborne fibers and total particulate matter. Fiber levels reported ranged<br />

from 0.01 to 1.4 f cm −3 , compared with 0.1–0.3 f cm −3 for glass wool.<br />

Total particulate matter sample results ranged from 0.05 to 23.6 mg m −3 in<br />

the mineral wool facilities and 0.09–8.48 mg m −3 for glass wool (Esmen et<br />

al., 1980).<br />

Comparison <strong>of</strong> Human MMVF exposures used in the<br />

recent rat chronic inhalation studies<br />

When using animal inhalation studies for assessment <strong>of</strong> potential risk to<br />

human health <strong>of</strong> airborne fibers, it is critical to demonstrate that the<br />

characteristics and concentrations <strong>of</strong> the experimental fiber aerosols are<br />

comparable with those in human exposure situations. It is also important<br />

for risk assessment that the actual target organ dose in the animal model<br />

reach or exceed that found in exposed humans. To illustrate, consider<br />

levels <strong>of</strong> fiber glass published in a number <strong>of</strong> recent reports. A qualitative


Table 7.6 Representative airborne levels <strong>of</strong> fiber glass in workplace and rat<br />

inhalation study<br />

a Outdoor data from Tiesler and Draeger (1994). Product-related fibers counted<br />

using NIOSH A Rules.<br />

b Data from Jacob et al., (1993). Airborne levels resulting from manufacturing<br />

operations using FG insulation.<br />

c Data from Lees et al., (1993a). Installation <strong>of</strong> residential insulation.<br />

d Jacob et al. (1992) reported that levels returned to background within hours after<br />

Batt installation.<br />

All other data are averages from the various studies herein cited.<br />

and quantitative comparison was made <strong>of</strong> the aerosol and lung fibers in the<br />

rat inhalation study with those in various human exposure situations<br />

(Hesterberg and Hart, 1994). A comparison <strong>of</strong> the reported aerosol fiber<br />

levels in various human settings with those used in the rat inhalation study<br />

is shown in Table 7.6. FG levels in the rat aerosol were more than five<br />

orders <strong>of</strong> magnitude higher than the reported level for outdoor air, and at<br />

least three orders <strong>of</strong> magnitude higher than for average airborne levels for<br />

many occupational settings (e.g. over 2000fold higher than FG batt<br />

installation). The rat aerosol was 75-fold more concentrated than the<br />

highest reported average TWA for airborne fiber levels in an occupational<br />

setting, i.e. blowing installation <strong>of</strong> unbound fiber glass (the potential for<br />

higher airborne levels has been recognized for some time, and<br />

recommended work practices call for the use <strong>of</strong> respirators in such<br />

circumstances). Despite the range in products and occupational settings,<br />

fiber dimensions in most <strong>of</strong> the human exposures examined were fairly<br />

similar to those found in the rat inhalation study aerosol (Hesterberg and<br />

Hart, 1994). The fiber dimensions <strong>of</strong> aerosolized rock and slag wool<br />

collected from workplace air during the installation <strong>of</strong> batts or blowing <strong>of</strong><br />

loose fibers have similar mean diameters to that <strong>of</strong> fiber glass (1.0–1.6<br />

µm). However, the mean lengths appear to be greater (30–50 µm) than for<br />

most workplace samples <strong>of</strong> fiber glass.<br />

Hesterberg and Hart (1994) also compared the lung burdens <strong>of</strong> rats<br />

exposed in the recent fiber glass inhalation study in rats with lung burdens<br />

found in workers involved in MMMF (primarily FG) manufacturing<br />

(McDonald et al. 1990). As shown in Table 7.7, rat fiber glass lung<br />

burdens vastly exceeded that <strong>of</strong> the workers reported by McDonald et al.,


which was not significantly elevated above reference levels. Fibers per mg<br />

dry lung for the rat after lifetime exposure was >4000-fold higher than for<br />

the fiber glass worker, average exposure 11 years (the average time from<br />

last employment in MMMF manufacturing and death was 12 years). Lung<br />

fiber dimensions in the rat study were comparable to those <strong>of</strong> fibers<br />

recovered from the lung tissue <strong>of</strong> fiber glass manufacturing workers. From<br />

these comparisons, it can be concluded that the exposure levels used in the<br />

recent rat inhalation studies unequivocally achieved the goal <strong>of</strong> the studies<br />

to exceed human exposures by several orders <strong>of</strong> magnitude.<br />

Summary and conclusions<br />

MMVFs are among the most studied commercial products due to their<br />

widespread use and the concern for potential health effects <strong>of</strong> respirable<br />

fibers. In recent animal inhalation studies RCF produced lung fibrosis,<br />

mesotheliomas, and significant increases in lung tumors. However, it is<br />

believed that any potential cancer risk from RCF exposure can be<br />

minimized, if not eliminated, because <strong>of</strong> the small number <strong>of</strong> workers<br />

exposed and the ability to use respiratory protection and engineering<br />

controls to limit worker exposure. Both human and animal inhalation<br />

studies have shown no association between fiber glass exposure and<br />

disease. Although high exposure levels <strong>of</strong> rock wool (several orders <strong>of</strong><br />

magnitude higher than most reported workplace exposures) produced<br />

minimal lung fibrosis in rats, no mesotheliomas and no significant increase<br />

in lung tumors were observed. Slag wool produced no fibrosis or increase<br />

in tumors in the animal studies. The cohort mortality studies <strong>of</strong> rock wool<br />

and slag wool workers have also provided no clear dose-response<br />

relationship with fiber exposure.<br />

Results from the combined animal inhalation studies showed that<br />

differences in lung fiber burdens and lung clearance rates could not explain<br />

the differences observed in the toxicologic effects <strong>of</strong> MMVFs. These<br />

findings clearly indicate that dose, dimension and durability (i.e. the<br />

persistence <strong>of</strong> fibers in the rat lung) are not the only determinants <strong>of</strong> fiber<br />

toxicity; chemical composition and the surface physicochemical properties<br />

<strong>of</strong> the fibers may also play an important role. Exposure levels from animal<br />

inhalation studies were at least three orders <strong>of</strong> magnitude higher than for<br />

average airborne levels reported for many occupational settings.

Table 7.7 Reported lung fiber levels from fiber glass workers and rat inhalation study<br />


a Lung fibers: for humans, NIOSH A rules; for rats, total fibers (all fibers length/diameter >3:1).<br />

b For humans, NIOSH A rules; for rats, WHO respirable fibers, comparable to A rules because there were no diameters >3 µm in<br />

rats.<br />

c Hesterberg et al., (1993b), Rat fiber exposure was 5 days week•1 , 6 h day•1 for lifetime (2 years).<br />

d McDonald et al., (1990). Negative controls had not worked with FG and were matched with each FG worker for age and<br />

locale.<br />

e Occupational exposures averaged 11 years, followed by average <strong>of</strong> 12 years without exposure prior to death.<br />

f 101 were FG workers; 11 were mineral wool workers.<br />

g Not reported.


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8<br />

Pulmonary Toxicity Studies with Man-made<br />

Organic Fibres: Preparation and Comparisons <strong>of</strong><br />

Size-separated Para-aramid with Chrysotile<br />

Asbestos Fibres<br />



1 DuPont Haskell Laboratory, Newark, DE, 2 Texas Tech<br />

Health Sciences Center Lubbock, TX<br />

Introduction<br />

This study was designed to compare the pulmonary toxic effects <strong>of</strong><br />

inhaled, size-separated preparations <strong>of</strong> chrysotile asbestos fibres with paraaramid<br />

fibrils at similar aerosol fibre concentrations. Chrysotile samples<br />

are known to have an abundance <strong>of</strong> short fibres, with mean lengths<br />

generally in the range <strong>of</strong> 2 µm. This is important to note because one <strong>of</strong> the<br />

critical factors influencing the pathogenesis <strong>of</strong> fibre-related lung disease is<br />

fibre dimension (Davis et al., 1986). As a consequence, attempts were made<br />

to selectively enhance the mean lengths <strong>of</strong> chrysotile asbestos fibres used in<br />

this inhalation toxicity study, in order to make reasonable comparisons<br />

between the two fibre-types.<br />

Methods<br />

General experimental design<br />

Groups <strong>of</strong> male Crl: CDBR rats (7–8 weeks old, Charles River Breeding<br />

Laboratories, Kingston, New York) were used to assess the pulmonary<br />

effects <strong>of</strong> 2-week inhalation exposures to size-separated preparations <strong>of</strong><br />

Kevlar ® para-aramid fibrils or chrysotile asbestos fibres. Animals were<br />

exposed 6 hr day −1 , 5 days week −1 for 2 weeks. For this study, Kevlar ® was<br />

utilized as a representative para-aramid fibril. The two commercial types <strong>of</strong><br />

para-aramid fibres are Twaron ® , made by Akzo, and Kevlar ® , made by<br />

DuPont. Following exposure, the lungs <strong>of</strong> p-aramid or chrysotile-exposed<br />

animals and age-matched sham controls were subsequently evaluated by<br />

bronchoalveolar lavage fluid analysis at 0 h, 5 days, 1 and 3 months<br />

postexposure. The lungs <strong>of</strong> additional animals were evaluated for<br />

biodurability, pulmonary clearance, pulmonary histopathologic lesions and


lung and mesothelial cell proliferation at 0 hrs, 5 days 1, 3, 6 and 12<br />

months postexposure.<br />

Fibre preparation and inhalation exposure<br />

Ultrafine Kevlar ® p-aramid fibrils were supplied by DuPont Fibres. A<br />

special preparation <strong>of</strong> respirable p-aramid fibrils which had been prepared<br />

for the 2-year inhalation study (Lee et al., 1988) was utilized for this study.<br />

Bulk Canadian chrysotile asbestos fibres were obtained from Mr John<br />

Addison <strong>of</strong> the Institute <strong>of</strong> Occupational Medicine in Edinburgh, Scotland.<br />

Attempts were made to size-separate the bulk fibre preparation (i.e.<br />

selectively enhance the percentages <strong>of</strong> long fibres while removing the short<br />

fibres) by placing the fibres in a rotating sieve shaker and sieving through a<br />

series <strong>of</strong> metal mesh screens. The fraction containing the longer fibres (and<br />

a number <strong>of</strong> short fibres) was collected and generated for inhalation<br />

studies; fibres were collected on a filter and dimensional analysis (i.e. length<br />

and diameter assessments) was performed using scanning electron<br />

microscopy. The results showed that this technique was partially successful<br />

as the median and mean lengths <strong>of</strong> fibres were increased from 3 and 5 µm,<br />

respectively, in the original bulk sample to 6 and 9 µm in the generated<br />

sample preparation. The median lengths and diameters <strong>of</strong> p-aramid fibrils<br />

used in the study were 9 µm and 0.3 µm, respectively.<br />

The methods for aerosol generation <strong>of</strong> p-aramid fibrils have previously<br />

been reported (Warheit et al., 1992). Final mean fibre concentrations for<br />

the p-aramid exposures were 772 and 419 f cm −3 .<br />

Aerosols <strong>of</strong> chrysotile asbestos fibres were generated in a similar<br />

manner, i.e. with a binfeeder and baffles, but without the microjet<br />

apparatus. Final mean fibre concentrations for the chrysotile asbestos<br />

exposures were 782 and 458 f cm −3 . Fibre lung burdens were quantified<br />

from digested lung tissue <strong>of</strong> animals sacrificed immediately after the end <strong>of</strong><br />

the 2-week exposure.<br />

Pulmonary lavage and biochemical assays on lavaged<br />

fluids<br />

Bronchoalveolar lavage procedures, cell quantification, and biochemical<br />

assays were conducted according to methods previously described (Warheit<br />

et al., 1984a, 1992). In addition, the methods for measuring lactate<br />

dehydrogenase (LDH), N-acetyl-β-glucosaminidase (NAG), and alkaline<br />

phosphatase (ALP) and protein in BAL fluids have been reported (Warheit<br />

et al., 1992).

Lung dissection, tissue preparation and pulmonary cell<br />

proliferation<br />

The lungs <strong>of</strong> rats exposed to p-aramid and chrysotile asbestos fibres for 2<br />

weeks were prepared for light microscopy by airway infusion using<br />

methods previously reported (Warheit et al., 1984b, 1991).<br />

Pulmonary cell proliferation experiments were designed to measure the<br />

effects <strong>of</strong> fibre inhalation exposure on terminal bronchiolar, proximal lung<br />

parenchymal (i.e. alveolar duct bifurcations and adjacent areas), subpleural<br />

and visceral pleural, and mesothelial cell turnover in rats following 2-week<br />

exposures. Groups <strong>of</strong> sham and fibre-exposed rats were given a 2-h pulse<br />

immediately after exposures, as well as 5 days, 1, 3, 6 and 12 months (still<br />

in progress) postexposure with an intraperitoneal injection <strong>of</strong> 5-bromo-2′deoxy-uridine<br />

(BrdU) dissolved in a 0.5N sodium bicarbonate buffer<br />

solution at a dose <strong>of</strong> 100 mg kg −1 body weight as previously described<br />

(Warheit et al., 1992). In addition, sections <strong>of</strong> duodenum served as a<br />

positive control. For each treatment group, there were immunostained<br />

nuclei in airways (i.e. terminal bronchiolar epithelial cells), lung parenchyma<br />

(i.e. epithelial, interstitial cells or macrophages), subpleura and visceral<br />

pleura, and mesothelial cells. All regions were counted by light microscopy<br />

at ×1000 magnification. Statistics were carried out using a two-tailed<br />

Students t test on a Micros<strong>of</strong>t Excel s<strong>of</strong>tware program.<br />

Fibre recovery from lung tissue<br />

Para-aramid fibrils were recovered from the lungs <strong>of</strong> exposed rats using a<br />

diluted 1.3% hypochlorite (Clorox bleach) solution. The results <strong>of</strong><br />

validation studies in our laboratory demonstrated that the dilute Clorox<br />

solution (10 min digestion) was more effective in digesting lung tissue than<br />

the KOH method that we had previously reported (Warheit et al., 1992).<br />

Chrysotile asbestos fibres were recovered from the lungs <strong>of</strong> exposed rats<br />

by incubating the lung tissue with a 5.25% hypochlorite solution for 3 h.<br />

Subsequently, the filters containing fibres recovered from lung tissue were<br />

mounted and prepared for phase-contrast light microscopy (for counting)<br />

and for scanning electron microscopy (for fibre dimensional analysis),<br />

according to methods previously described (Warheit et al., 1992).<br />

Results<br />

D.B.WARHEIT ET AL. 119<br />

Size-separation methods for chrysotile asbestos fibres<br />

The results from size-separation attempts showed that there was a shift in<br />

the distribution <strong>of</strong> fibre lengths from shorter fibres to longer fibres<br />

(Figures 8.1(A)– (C)). Count median lengths <strong>of</strong> chrysotile asbestos fibres


were increased from 3 µm in the original generated sample to 6 µm in the<br />

size-separated sample. In comparison to the chrysotile asbestos sample,<br />

there was a significantly greater proportion <strong>of</strong> long p-aramid fibrils which<br />

were used in the inhalation study with median lengths >9 µm.<br />

Lung burden analysis<br />

Although the aerosol fibre concentrations were similar throughout the<br />

study (p-aramid high conc.=772 f cm −3 , chrysotile high conc.=782 f cm −3 ;<br />

p-aramid low conc.=419 f cm −3 , chrysotile low conc.=458 f cm −3 ),<br />

measurement <strong>of</strong> lung fibre burdens from digested lung tissue at time 0 (i.e.<br />

immediately after exposure) demonstrated a substantial difference in lung<br />

burden between the two fibre-types as measured by phase contrast optical<br />

microscopy (PCOM). The mean lung fibre (>5 µm) burden from 3 rats/<br />

dose group exposed to chrysotile asbestos was 3.7×10 7 (±7.4×10 6 ) fibres/<br />

lung for the high dose group and 1.3×10 7 (±4×10 6 ) fibres/lung for the low<br />

dose group. In contrast, the mean lung fibre burden from 3 rats/dose group<br />

exposed to para-aramid fibres was 7.6×10 7 (±1.9×10 7 ) fibres per lung for<br />

the high dose group and 4.8×10 7 ( ±2.1×10 7 ) fibres/lung for the low dose<br />

group. In addition, the count median length <strong>of</strong> chrysotile fibres recovered<br />

from the lungs <strong>of</strong> exposed animals immediately after 2-week exposure was<br />

3.5 µm, while the count median diameter was 0.15 µm. In contrast, the<br />

count median length <strong>of</strong> para-aramid fibres recovered from the lungs <strong>of</strong><br />

exposed animals immediately after 2-week exposure was 8.6 µm, while the<br />

count median diameter was 0.3 µm (Figure 8.2(A) and (B); numerical data<br />

not shown). These data indicate that our attempts to size-separate<br />

Canadian chrysotile fibres were only partially successful. The lung burden<br />

data also suggest that comparisons <strong>of</strong> the effects <strong>of</strong> chrysotile vs paraaramid<br />

at high and low doses are difficult to make since the doses were not<br />

equivalent.<br />

Bronchoalveolar lavage data<br />

Two-week exposures to p-aramid fibrils or chrysotile asbestos fibres<br />

produced transient pulmonary inflammatory responses as measured by<br />

bronchoalveolar lavage fluid analysis (see Table 8.1).<br />

Light microscopic histopathology<br />

Exposures to p-aramid and chrysotile were associated with minimal to mild<br />

centriacinar inflammation and fibrosis (increased trichrome staining)<br />

immediately after and 5 days after 2-week exposures. Lesions were slightly<br />

more prominent in p-aramid-exposed rats due to increased inflammation.<br />

Lesions were less severe at 1 month and essentially resolved at 6 months

D.B.WARHEIT ET AL. 121<br />

Figure 8.1 (A) Chrysotile asbestos lengths—original generated sample for 4 different<br />

experiments. The graph depicts the fibre length distributions as assessed by<br />

scanning electron microscopy from four aerosol exposures prior to attempts to size<br />

separate the fibres. Fifty percent <strong>of</strong> the fibres from all four groups are less than 3–4<br />

µm. (B) Distributions <strong>of</strong> size-separated chrysotile asbestos lengths used in the<br />

inhalation study from the high-dose


Figure 8.1 Continued<br />

exposures and (C) the low-dose exposure groups. A casual glance at the two graphs<br />

B and C indicates that some success was attained in increasing the mean lengths in<br />

the aerosol <strong>of</strong> the generated chrysotile asbestos sample.<br />

with only occasional centriacinar regions having slight, fibril-associated<br />

thickening <strong>of</strong> alveolar duct bifurcations. At 1 year postexposure, the lungs<br />

in p-aramid exposed rats were similar to controls. The 1-year chrysotileexposed<br />

animals are still in recovery.<br />

Pulmonary cell proliferation<br />

In chrysotile asbestos-exposed rats, substantial increases compared to<br />

controls in pulmonary cell proliferation indices were measured on terminal<br />

bronchiolar, parenchymal, visceral pleural/subpleural and mesothelial<br />

surfaces, and many <strong>of</strong> these effects were sustained through 3 months<br />

postexposure. These data demonstrate that 2-week chrysotile exposures<br />

produced a prolonged proliferative response in airway, alveolar and<br />

subpleural cells, as evidenced by the sustained effect through 3 months<br />

postexposure (Table 8.2).<br />

Pulmonary cell proliferation studies demonstrated that 2-week exposures<br />

to the high dose <strong>of</strong> p-aramid fibrils produced a transient increase in<br />

terminal bronchiolar and visceral pleural/subpleural cell labeling responses.<br />

No increases in lung parenchymal, or subpleural cell labeling indices were<br />

mea sured at any time period relative to sham controls. In addition, no

D.B.WARHEIT ET AL. 123<br />

Figure 8.2 (A) Scanning electron microscopy (SEM) micrograph <strong>of</strong> an aerosol filter<br />

containing a mixture <strong>of</strong> long and short chrysotile asbestos fibres (arrows). (B) An<br />

SEM micrograph <strong>of</strong> fibres recovered from the lung <strong>of</strong> a rat 3 months after 2-week<br />

chrysotile exposures. Note that most <strong>of</strong> the fibres are long (arrows), indicating that<br />

the long chrysotile asbestos fibres were retained in the lung while the shorter fibres<br />

were cleared from the respiratory tract.<br />

increases in cell labeling indices were measured in animals exposed to a<br />

lower dose <strong>of</strong> p-aramid fibrils at any postexposure time period (Table 8.2).


Table 8.1 Pulmonary inflammation and fibre biodurability in the lungs <strong>of</strong><br />

chrysotile asbestos and p-aramid-exposed rats<br />

0 h=immediately after exposure; 5 D=5 days; 1 M=1 month; 3 M=3 months; 6<br />

M=6 months;<br />

ND=not determined.<br />

Lung digestion/biodurability studies<br />

Preliminary dimensional analysis studies demonstrated that median lengths<br />

<strong>of</strong> fibres recovered from digested asbestos-exposed lung tissue were<br />

increased over time suggesting that short asbestos fibres were selectively<br />

cleared from the lungs, with apparent insignificant or pulmonary clearance<br />

and greater durability/retention <strong>of</strong> long fibres (Table 8.1).<br />

Preliminary studies with p-aramid fibrils recovered from the lungs <strong>of</strong><br />

exposed rats are consistent with earlier data suggesting biodegradability <strong>of</strong><br />

inhaled p-aramid fibrils (Warheit et al., 1992; Kelly et al., 1993)<br />

(Table 8.1). These data also are in agreement with the results <strong>of</strong> a current<br />

interim report authored by the Institute <strong>of</strong> Occupational Medicine in<br />

Edinburgh, Scotland. In addition, as previously reported (Warheit et al.,<br />

1992), a transient increase in fibre numbers at early postexposure time<br />

periods was measured following cessation <strong>of</strong> exposure. These results<br />

indicate that the increase in p-aramid fibres is due to fibre shortening and as<br />

a consequence, increased numbers <strong>of</strong> shorter fibres. This is accounted for<br />

by a substantial reduction in the median lengths <strong>of</strong> recovered fibres<br />

concomitant with only a slight decrease in fibre diameter.

Table 8.2 Cell proliferation effects in chrysotile asbestos and p-Aramid-exposed<br />

rats<br />

a p


The BrdU pulmonary cell labeling results demonstrating sustained<br />

proliferative effects in chrysotile-exposed rats presented here are consistent<br />

with findings from several other investigators (Brody and Overby, 1989;<br />

McGavran et al., 1990; Coin et al., 1992a). In studies by Brody and<br />

Overby (1989), acute inhalation exposures to chrysotile asbestos fibres<br />

produced a biphasic cell labeling response in the lungs <strong>of</strong> exposed rats and<br />

mice. This was characterized by dramatic increases in epithelial cell DNA<br />

synthesis, followed several days later by enhanced labeling <strong>of</strong> interstitial<br />

cells. In follow-up studies, a 3 day exposure prolonged the duration <strong>of</strong><br />

increased cell labeling (Coin et al., 1992b). In another study, Coin et al.,<br />

(1991) reported that a 5-h exposure to chrysotile fibres in mice produced<br />

substantial increases in mesothelial and subpleural cell labeling indices at 2<br />

and 8 days postexposure.<br />

The finding <strong>of</strong> sustained subpleural and mesothelial cell proliferation in<br />

chrysotile-exposed rats was unexpected and raises the issue regarding the<br />

association <strong>of</strong> chrysotile with the development <strong>of</strong> mesothelioma. In this<br />

regard, inhalation <strong>of</strong> chrysotile asbestos fibres produced mesotheliomas in<br />

exposed rats (Wagner et al., 1974; Davis and Jones, 1988).<br />

The biodurability data reported here demonstrating retention or reduced<br />

clearance <strong>of</strong> long chrysotile fibres are consistent with the results <strong>of</strong><br />

previous studies by Roggli and Brody (1984) and Bellmann et al., (1986,<br />

1987). In contrast to the enhanced biodurability <strong>of</strong> chrysotile asbestos<br />

fibres, the results with p-aramid fibres suggest that the fibrils undergo<br />

biodegradability in the lungs <strong>of</strong> exposed rats. These findings confirm our<br />

earlier studies (Warheit et al., 1992) and are in concordance with the<br />

results <strong>of</strong> Kelly et al. (1993) and the recent findings <strong>of</strong> the IOM.<br />

In conclusion, size separation techniques for chrysotile asbestos fibres<br />

were partially successful in increasing median lengths from 3 µm to 6 µm.<br />

Histopathological studies demonstrated that both p-aramid and chrysotile<br />

produced a minimal to mild inflammatory response which produced<br />

thickening <strong>of</strong> the alveolar duct bifurcations. These effects peaked at 1<br />

month postexposure and were essentially reversible by 6 months<br />

postexposure.<br />

Pulmonary cell labeling studies demonstrated substantial increases in<br />

lung parenchymal, airway, pleural/subpleural, and mesothelial cell<br />

proliferation effects following chrysotile exposures, suggesting that<br />

chrysotile produces a potent proliferative response in the airways, lung<br />

parenchyma, and subpleural/ pleural regions. In contrast, p-aramid<br />

exposures produced only transient effects in airway and subpleural regions.<br />

Fibre biopersistence/durability results thus far indicate that the long<br />

chrysotile fibres are retained in the lung or cleared at a slow rate. In<br />

contrast, p-aramid fibres have low biodurability in the lungs <strong>of</strong> exposed<br />

animals. In this regard, median lengths <strong>of</strong> chrysotile fibres recovered from

exposed lung tissue were increased over time, while median lengths <strong>of</strong> paramid<br />

fibrils were decreased over time.<br />

It is concluded that the proliferative effects and enhanced biodurability<br />

<strong>of</strong> chrysotile that are associated with the induction <strong>of</strong> chronic disease do<br />

not occur with p-aramid fibrils. Therefore, inhalation <strong>of</strong> chrysotile asbestos<br />

fibres is likely to produce greater long-term pulmonary toxic effects in<br />

comparison to para-aramid fibrils.<br />

Acknowledgments<br />

This study was sponsored by the DuPont Co. and Akzo Nobel Corp.<br />

References<br />

D.B.WARHEIT ET AL. 127<br />

BELLMANN, B., KONIG, H., MUHLE, H. and POTT, F., 1986, Chemical<br />

durability <strong>of</strong> asbestos and <strong>of</strong> man-made mineral fibres in vivo, J. Aerosol Sci.,<br />

17, 341–5.<br />


SPURNY, K., 1987, Persistence <strong>of</strong> man-made mineral fibers (MMMF) and<br />

asbestos in rat lungs, Ann. Occup. Hyg., 31(4B), 693–709.<br />

BRODY, A.R. and OVERBY, L.H., 1989, Incorporation <strong>of</strong> tritiated thymidine by<br />

epithelial and interstitial cells in bronchiolar-alveolar regions <strong>of</strong> asbestosexposed<br />

rats, Am. J. Pathol., 134, 133–40.<br />

COIN, P.G., MOORE, L.B., ROGGLI, V. and BRODY, A.R., 1991, Pleural<br />

incorporation <strong>of</strong> 3 H-TdR after inhalation <strong>of</strong> chrysotile asbestos in the mouse,<br />

Am. Rev. Respir. Dis., 143, A604.<br />

COIN, P.G., ROGGLI, V.L. and BRODY, A.R., 1992a, Deposition, clearance and<br />

translocation <strong>of</strong> chrysotile asbestos from peripheral and central regions <strong>of</strong> the<br />

rat lung, Environ, Res., 58, 97–116.<br />

COIN, P.G., ROGGLI, V. and BRODY, A.R., 1992b, Pulmonary fibrogenesis and<br />

BRDU incorporation after three consecutive inhalation exposures to chrysotile<br />

asbestos, Am. Rev. Respir. Dis., 145, A328.<br />

DAVIS, J.M.G. and JONES, A.D., 1988, Comparisons <strong>of</strong> the pathogenicity <strong>of</strong> long<br />

and short fibres <strong>of</strong> chrysotile asbestos in rats, Br. J. Exp. Pathol., 69, 717–37.<br />

DAVIS, J.M.G., ADDISON, J., BOLTON, R.E., DONALDSON, K. et al., 1986,<br />

The pathogenicity <strong>of</strong> long versus short fibre samples <strong>of</strong> amosite asbestos<br />

administered to rats by inhalation or intraperitoneal injection, Br. J. Exp.<br />

Pathol, 67, 415–30.<br />

KELLY, D.P., MERRIMAN, E.A., KENNEDY, G.L.JR. and LEE, K.P., 1993,<br />

Deposition, clearance, and shortening <strong>of</strong> Kevlar para-aramid fibrils in acute,<br />

subchronic, and chronic inhalation studies in rats, Fundam. Appl. Toxicol, 21,<br />

345–54.<br />

LEE, K.P., KELLY, D.P., O’NEAL, F.O., STADLER, J.C. and KENNEDY, G.<br />

L.JR, 1988, Lung response to ultrafine Kevlar aramid synthetic fibrils<br />

following 2-year inhalation exposure in rats, Fundam. Appl. Toxicol., 11, 1–<br />



McGAVRAN, P.D., BUTTERICK, C.J. and BRODY, A.R., 1990, Tritiated<br />

thymidine incorporation and the development <strong>of</strong> an interstitial lesion in the<br />

bronchiolar alveolar regions <strong>of</strong> the lungs <strong>of</strong> normal and complement deficient<br />

mice after inhalation <strong>of</strong> chrysotile asbestos, J. Environ. Pathol. Toxicol.<br />

Oncol, 9, 377–92.<br />

ROGGLI, V.L. and BRODY, A.R., 1984, Changes in numbers and dimensions <strong>of</strong><br />

chrysotile asbestos fibers in lungs <strong>of</strong> rats following short-term exposure, Exp.<br />

Lung Res., 7, 133–47.<br />

WAGNER, J.C., BERRY, G., SKIDMORE, J.W. and TIMBRELL, V., 1974, The<br />

effects <strong>of</strong> the inhalation <strong>of</strong> asbestos in rats, Br. J. Cancer, 29, 252–70.<br />

WARHEIT, D.B., HILL, L.H. and BRODY, A.R., 1984a, Surface morphology and<br />

correlated phagocytic capacity <strong>of</strong> pulmonary macrophages lavaged from the<br />

lungs <strong>of</strong> rats, Exp. Lung Res., 6, 71–82.<br />

WARHEIT, D.B., CHANG, L.Y., HILL, L.H., HOOK, G.E.R., CRAPO, J.D. and<br />

BRODY, A.R., 1984b, Pulmonary macrophage accumulation and<br />

asbestosinduced lesions at sites <strong>of</strong> fiber deposition, Am. Rev. Respir. Dis., 129,<br />

301.<br />


1991, Development <strong>of</strong> a short-term inhalation bioassay to assess pulmonary<br />

toxicity <strong>of</strong> inhaled particles: Comparisons <strong>of</strong> pulmonary responses to carbonyl<br />

iron and silica, Toxicol Appl. Pharmacol., 107, 350–68.<br />

WARHEIT, D.B., KELLAR, K.A. and HARTSKY, M.A., 1992, Pulmonary cellular<br />

effects in rats following aerosol exposures to ultrafine Kevlar® aramid fibrils:<br />

evidence for biodegradability <strong>of</strong> inhaled fibrils, Toxicol. Appl. Pharmacol,<br />

116, 225– 39.

9<br />

Pulmonary Hyperreactivity to <strong>Industrial</strong><br />

Pollutants<br />


Bayer AG, Wuppertal<br />

Introduction<br />

Environmental agents, such as ozone, nitrogen dioxide, formaldehyde, and<br />

sulfur dioxide; occupational pollutants, including natural dusts (grain, red<br />

cedar, animal dander), irritant fumes or vapors, and organic acid<br />

anhydrides, reactive dyes, or (di)isocyanates can cause increases in airway<br />

reactivity. Airway hyperreactivity is defined as an exaggerated acute<br />

obstructive response <strong>of</strong> the airways to one or more nonspecific stimuli. The<br />

incriminated etiologic low-molecular-weight agents all share a common<br />

toxicological characteristic <strong>of</strong> being irritant in nature. In some cases, the<br />

agents are present as a gas, in others the inciting agent is an aerosol. As yet<br />

it is not clear, for instance, whether induced airway hyperreactivity is a<br />

dose-effect phenomenon and whether a brief high level exposure causes<br />

more prolonged or intense airways response. While the illness clinically<br />

simulates bronchial asthma and is associated with airway hyperreactivity,<br />

it is considered to be different from typical occupational asthma because <strong>of</strong><br />

its rapid onset, specific relationship to a single environmental exposure,<br />

and no apparent preexisting period <strong>of</strong> sensitization to occur with the<br />

apparent lack <strong>of</strong> an allergic or immunologic etiology. Hence, this illness is<br />

termed reactive airways dysfunction syndrome, or RADS, because the<br />

characteristic finding is hyperreactivity <strong>of</strong> the airways (Brooks et al.,<br />

1985). Mechanisms to explain the development <strong>of</strong> RADS focus on the<br />

toxic effects <strong>of</strong> the irritant exposure on the airways. How this increased<br />

bronchial responsiveness is precisely triggered, amplified, sustained and<br />

how it relates to inflammatory events remains, to a certain extent,<br />

incompletely elucidated (Kay, 1991).<br />

A common pathologie accompaniment or cause <strong>of</strong> increased airway<br />

hyper-responsiveness is pulmonary inflammation. It is suggested that this<br />

inflammation is responsible for the change in histamine or cholinergic<br />

agonist responsiveness. Because subepithelial irritant receptors are<br />

superficial in location, they could be affected by an extensive bronchial<br />

inflammatory response which might occur after heavy irritant exposure.


Subsequent re-epithelialization and probable reinervation <strong>of</strong> bronchial<br />

mucosa might drastically alter the threshold <strong>of</strong> the receptors and cause<br />

airways hyperreactivity. It has been hypothesized that damage to airway<br />

epithelium by irritant chemicals could decrease the threshold <strong>of</strong> sensory<br />

endings within the mucosa, resulting in increased afferent and efferent vagal<br />

activity. Airway mucosal inflammation, activation <strong>of</strong> airway afferent<br />

nerves, and the release <strong>of</strong> low-molecular-weight neuropeptides as<br />

mediators <strong>of</strong> inflammation are known to affect the tonus <strong>of</strong> the airway<br />

smooth muscle and may play a crucial role in the acute increase in airway<br />

hyperresponsiveness occurring after exposure to irritant or inflammatory<br />

stimuli. Additionally, inflammatory mediators may further attract and<br />

activate inflammatory cells, which themselves release a whole array <strong>of</strong><br />

chemotactic and cytotoxic mediators that serve to perpetuate and amplify<br />

the inflammatory response. This complex interaction <strong>of</strong> different factors<br />

may result in epithelial desquamation, mucus gland hyperplasia, smooth<br />

muscle hypertrophy, and eventually render the airways hyperreactive to<br />

specific as well as nonspecific stimuli.<br />

Increased bronchial irritability, or hyperresponsiveness, to a wide variety<br />

<strong>of</strong> chemical agents and physical stimuli is also a major characteristic<br />

feature <strong>of</strong> bronchial asthma and the reactive airways dysfunction syndrome<br />

might clinically be indistinguishable from the asthma syndrome. Also for<br />

the latter, particular attention has been placed on the role <strong>of</strong> inflammation<br />

mediated influx <strong>of</strong> cells, mediator release and the interaction <strong>of</strong> irritant<br />

induced neurogenic and inflammatory factors. Neural control <strong>of</strong> airway<br />

caliber is far from being simple and it is likely to contribute to airway<br />

narrowing and bronchial hyper-responsiveness. Myelinated and<br />

nonmyelinated nerve fibers (C fibers) are involved in the sensory irritation<br />

response and their stimulation may result in release <strong>of</strong> specific<br />

neuropeptides, known to be potent releasers <strong>of</strong> mediators from airway<br />

mast cells (Barnes et al., 1991a, b; Nielsen, 1991). Specific neuropeptides<br />

are also known to attract eosinophils which can be stimulated to release<br />

cytotoxic mediators that may exacerbate these pseudoallergic-like<br />

responses even further. Experimental and clinical studies have intimated<br />

that there is reason to suspect that acute exposure to brief high-level<br />

concentrations <strong>of</strong> asthmagenic chemicals and the development <strong>of</strong> increased<br />

airway hyperresponsiveness are associated. Thus, it could be assumed that<br />

specific mast cell sensitization—in combination with neurogenic stimuli—<br />

amplify the inflammatory process and airway hyperresponsiveness. The<br />

corresponding increase in vagal activity would increase reflex release <strong>of</strong><br />

acetylcholine and, correspondingly, may enhance airway responsiveness<br />

following the exogenous administration <strong>of</strong> cholinergic agents.<br />

Animal models <strong>of</strong> airway inflammation might allow us to investigate this<br />

relationship further. Models <strong>of</strong> allergic pulmonary inflammation have been<br />

developed in various animal species (Kips et al., 1992), using different

method- ological approaches. In toxicology, the guinea-pig has been used<br />

for decades in order to evaluate the skin sensitizing properties <strong>of</strong> chemicals<br />

and proteins and has also been able to reproduce immediate-onset<br />

pulmonary hypersensitivity responses following inhalation <strong>of</strong> chemical<br />

haptens, their protein-conjugates or antigens. This animal model has<br />

therefore been used to disclose principles governing both the development<br />

<strong>of</strong> pulmonary hypersensitivity and airway hyperreactivity. Due to the<br />

guinea-pig’s abundant amount <strong>of</strong> smooth bronchial musculature, it is used<br />

as a physiologic elicitation model that reproduces bronchospasm upon<br />

challenge to specific or nonspecific stimuli. Other animal models designed<br />

to display many <strong>of</strong> the chronic features <strong>of</strong> hypersensitivity lung diseases<br />

characteristic <strong>of</strong> occupational asthma focus more on the induction <strong>of</strong><br />

airway inflammation, the basic prerequisite for airway hyperreactivity. It<br />

should be noted, however, that the induction <strong>of</strong> asthma in the rat model,<br />

for example, commonly requires more aggressive protocols and more<br />

elaborate techniques to classify responses when compared with the guineapig<br />

elicitation model (vide infra).<br />

The guinea-pig model<br />

J.PAULUHN 131<br />

To date, practically all such models have relied upon the use <strong>of</strong> the guineapig,<br />

a species known to be sensitive for agents inducing<br />

bronchoconstriction and in which respiratory function and respiratory<br />

hypersensitivity can be measured readily. In addition, guinea-pigs are easy<br />

to handle, relatively inexpensive, and produce consistent<br />

bronchoconstrictive reactions. The models have utilized various modes <strong>of</strong><br />

hapten or antigen administration and methods for detecting sensitization,<br />

It has been shown that guinea-pigs sensitized by inhalation exposure to<br />

either a free or a protein-bound chemical can be induced to exhibit changes<br />

in respiratory patterns following inhalation challenge with the same<br />

chemical in the free or in the form <strong>of</strong> its hapten-protein conjugate. In the<br />

guinea-pig no adjuvant is needed for successful lung sensitization. More<br />

recently it has been found that changes in sensitive respiratory parameters<br />

can also be provoked in dermally sensitized guinea-pigs by inhalation<br />

challenge with the free chemical or the hapten-protein conjugate (Botham et<br />

al., 1988; Pauluhn and Eben, 1991; Hayes et al., 1992). In attempting to<br />

derive an animal model that permits the identification <strong>of</strong> asthmagenic lowmolecularweight<br />

chemicals without the presence <strong>of</strong> overriding effects<br />

caused by toxic (irritant) airway inflammation the intradermal route <strong>of</strong><br />

induction appears to be preferable. This route <strong>of</strong> induction also minimizes<br />

the risk <strong>of</strong> potential confounding effects attributable to irritant-induced<br />

nonspecific reactive airways dysfunction as a result <strong>of</strong> previous inhalation<br />

exposures (Briatico-Vangosa et al., 1993).


For challenge exposures it appears to be advantageous to use the free<br />

chemical in slightly irritant concentrations rather than the proteinconjugate<br />

<strong>of</strong> the hapten. It is believed that the in vitro synthesis <strong>of</strong> the<br />

hapten-protein conjugate may not necessarily result in immunologically<br />

identical conjugates when compared with those produced under in vivo<br />

conditions. Also standardized procedures to synthesize and characterize<br />

hapten-protein conjugates <strong>of</strong> multifunctional, highly reactive chemicals are<br />

not yet established. On the other hand, an essential prerequisite for<br />

challenge exposures with the free chemical is the evaluation <strong>of</strong> the irritant<br />

threshold concentration <strong>of</strong> the hapten under investigation. The importance<br />

<strong>of</strong> concentration in distinguishing irritation from sensitization cannot be<br />

overstated and is one <strong>of</strong> the most critical determinants <strong>of</strong> this animal<br />

model.<br />

For volatile, irritant haptens the characteristic feature <strong>of</strong> upper<br />

respiratory tract irritation is the reflexively induced decrease in respiratory<br />

rate which is a common finding in laboratory rodents (Figure 9.1).<br />

Consistent with this approach, naive mice, rats and guinea-pigs were<br />

exposed for 45 min to slightly irritant concentrations <strong>of</strong> phenyl isocyanate<br />

(PI). As evident from Figure 9.1, the exposure to ca. 5 mg PI m −3 air<br />

provoked a decrease in respiratory rate <strong>of</strong> approximately 25–45%. The<br />

observation that remarkable differences in response patterns between mice,<br />

rats and guinea-pigs did not occur demon strate that irritant threshold<br />

concentrations obtained in mice may also be valid for guinea-pigs. Mainly<br />

for volatile chemicals attempts have been made to establish methods for the<br />

measurement and analysis <strong>of</strong> the irritant-induced changes in respiratory<br />

pattern in mice (Vijayaraghavan et al., 1993) and to understand the<br />

mechanisms <strong>of</strong> the irritant receptor stimulation (Nielsen, 1991). For<br />

volatile irritant haptens, such as PI, an unequivocal respiratory<br />

hypersensitivity response is characterized by a shallow rapid breathing<br />

pattern, i.e. a response opposite to that occurring as a result <strong>of</strong> upper<br />

respiratory tract irritation. For volatile irritant haptens this type <strong>of</strong><br />

breathing pattern, however, can only be obtained when using the proteinconjugate<br />

<strong>of</strong> the hapten.<br />

The interpretation <strong>of</strong> changes in respiratory pattern induced by irritant<br />

particulates is less predictable because <strong>of</strong> the size-dependent deposition <strong>of</strong><br />

particles within the respiratory tract. Irritant aerosols that evoke bronchial<br />

or pulmonary irritation may produce changes similar to those occurring<br />

following immediate-onset responses. Therefore, the selection <strong>of</strong> adequate<br />

haptenchallenge concentrations as well as the measurement <strong>of</strong> several<br />

breathing parameters is <strong>of</strong> primary importance. For such chemicals,<br />

currently the relative effectiveness <strong>of</strong> the acute high-concentration<br />

inhalation (single inhalation exposure <strong>of</strong> 15 min) and the high-dose<br />

intradermal route for sensitization <strong>of</strong> guinea-pigs had been investigated<br />

(Pauluhn and Eben, 1991; Pauluhn and Mohr, 1994). The airway function

J.PAULUHN 133<br />

Figure 9.1 Time-response curves for respiratory rate from mice, rats and guineapigs<br />

during single 45-min exposures to appoximately 5 mg m −3 phenyl isocyanate.<br />

Data were normalized on pre-exposure values during 15-min air exposure. Data<br />

points for each concentration are the mean <strong>of</strong> four animals and were averaged for<br />

45 s.<br />

<strong>of</strong> conscious guinea-pigs that were sensitized to and challenged with 4,4′diphenylmethane-diisocyanate<br />

(MDI) aerosol or trimellitic anhydride<br />

(TMA) dust as well as their corresponding proteinconjugates was<br />

monitored plethysmographically. The airway hyper-responsiveness to<br />

subsequently increased inhaled acetylcholine (ACh) concentrations was<br />

assessed 1 day after the hapten challenge (Pauluhn, 1994). In most<br />

instances, selected morphological features <strong>of</strong> the airways (increased<br />

number <strong>of</strong> eosinophils in the bronchial mucosa and lung associated lymph<br />

nodes) were also taken into account.<br />

Collectively, it was noticed that elicitation <strong>of</strong> respiratory hypersensitivity<br />

is concentration-dependent and that challenge concentrations should<br />

slightly exceed the threshold concentration for irritation. The evaluation <strong>of</strong><br />

eosinophils in subepithelial tissues and lung associated lymph nodes<br />

appears to provide an important independent adjunct to measurements <strong>of</strong><br />

respiratory function. The combined assessment <strong>of</strong> specific pathologic<br />

features such as eosinophilic infiltration and the evaluations <strong>of</strong> several


breathing parameters upon acetylcholine and hapten or conjugate challenge<br />

significantly enhance the diagnostic sensitivity <strong>of</strong> the guinea-pig model.<br />

From studies using single, brief high level aerosol or dust exposures for the<br />

induction <strong>of</strong> animals it can be concluded that previous high level exposures<br />

evoke bronchial hyperresponsiveness upon challenge at lower hapten<br />

concentrations when compared with intradermally sensitized animals.<br />

However, guinea-pigs sensitized intradermally to the volatile PI<br />

demonstrated remarkable immediate-type respiratory reactions only upon<br />

challenge with the conjugate and not with slightly irritant concentrations<br />

<strong>of</strong> the free PI. To study if phenyl isocyanate is capable <strong>of</strong> inducing a<br />

reactive airway or an asthma like syndrome, the subsequently described rat<br />

model was used.<br />

The rat model<br />

This animal model focuses on the induction <strong>of</strong> airway inflammation which<br />

comprises most <strong>of</strong> the characteristic features <strong>of</strong> asthma. It has been stated<br />

that respiratory hypersensitivity should depend on two separate factors:<br />

first, the degree <strong>of</strong> allergic airways, and second, the sensitivity to<br />

bronchoconstrictive mediators. Increasing evidence suggests that the<br />

eosinophils play a critical role in the pathogenesis <strong>of</strong> asthma and <strong>of</strong> other<br />

non-allergic hyperresponsive airway diseases. For the induction <strong>of</strong> the<br />

asthmatic state male rats were exposed for 2 consecutive weeks by<br />

inhalation (5 h day −1 , 5 days week −1 ). The target concentrations <strong>of</strong> phenyl<br />

isocyanate were chosen on the basis <strong>of</strong> a single 45-min exposure study<br />

which suggested that approximately 1 mg m −3 air is the irritant threshold<br />

concentration for ‘any’ duration <strong>of</strong> exposure. The 2 week repeated<br />

inhalation study was designed to assess the functional, bio chemical and<br />

morphological signs <strong>of</strong> phenyl isocyanate induced lung disease and their<br />

regression during an observation period <strong>of</strong> approximately 2 months.<br />

The most characteristic features <strong>of</strong> asthma comprise an increased influx<br />

<strong>of</strong> eosinophilic granulocytes into the tissue <strong>of</strong> the airways, secretory cell<br />

hyper-plasia and metaplasia, smooth muscle hypertrophy and hyperplasia,<br />

epithelial desquamation, airway hyperresponsiveness, and eventually partial<br />

occlusion <strong>of</strong> the airway lumen with mucus and cellular debris. The<br />

formation <strong>of</strong> mucus plugs is a regular feature <strong>of</strong> asthma and accounts for<br />

most <strong>of</strong> the clinical, biochemical and physiological abnormalities.<br />

Histopathological evaluation <strong>of</strong> the respiratory tract indicated a<br />

bronchiolitis obliterans and smooth muscle hypertrophy in rats exposed to<br />

approximately 7 mg m −3 air, whereas only minimal effects were found<br />

following 4 mg m −3 air. Lung function measurements revealed that some rats<br />

were hyperresponsive to an ACh-stimulus. As shown in Figure 9.2, also the<br />

increase in shunt blood (Q s/Q t) anddecrease in forced expiratory flow rates<br />

(MMEF) as well as mucus products (sialomucins), polymorphonuclear

cells, including eosinophils, in the bronchoalveolar lavage fluid (BALF)<br />

were consistent with an asthma like syndrome. As evident from Figure 9.2,<br />

the changes observed in rats exposed to 7 mg m −3 air did not fully regress<br />

during an observation period <strong>of</strong> approximately 2 months.<br />

Conclusion<br />

J.PAULUHN 135<br />

Figure 9.2 Relative comparison <strong>of</strong> sensitive diagnostic parameters in rats exposed to<br />

either 0 (air), 1, 4 and 7 mg PI m −3 air for 2 consecutive weeks (6 h day −1 , 5 days<br />

week −1 ). The measurements were performed in weeks 3 and 9. Abbreviations:<br />

MMEF: Maximal mid-expiratory flow rate, Q s/Q t: venous admixture, PMN:<br />

polymorphonuclear cells, Eos: eosinophilic granulocytes, Sialomucins: total sialic<br />

acid (after hydrolysis).<br />

Experimental evidence suggests that changes within the respiratory tract<br />

leading to the reactive airway dysfunction syndrome and/or asthma are<br />

fully consistent with an inflammatory response involving tissue <strong>of</strong> direct<br />

contact. The toxicity <strong>of</strong> irritant chemicals known to induce such illness is<br />

highly focal, and the variability <strong>of</strong> response in different regions <strong>of</strong> the<br />

respiratory tract could be a result <strong>of</strong> the actual concentration <strong>of</strong> the<br />

toxicant reaching various airway levels. Determination <strong>of</strong> immunologic<br />

etiology is particularly important for chemical allergy since all recognized<br />

low-molecular-weight chemical sensitizers are also respiratory irritants and<br />

in sufficient concentrations can cause airway constriction by<br />

nonimmunological mechanisms. As shown by studies using phenyl


isocyanate, damage <strong>of</strong> the airways is characterized by a steep concentrationresponse<br />

curve. Based on acute 45-min exposure <strong>of</strong> rats the threshold for<br />

respiratory tract irritation is approximately 1 mg m −3 . Exposures equaling<br />

this concentration were tolerated without exposure-related effects, whether<br />

exposure occurred singly for 45-min or repeatedly for 2 weeks. Marginal<br />

effects were observed at 4 mg m −3 , all effects, including mortality, were<br />

produced at 7 mg m −3 . This demonstrates that selection <strong>of</strong> appropriate<br />

exposure concentrations appears to be most critical in the rat model. The<br />

assessment <strong>of</strong> diagnostic sensitivity <strong>of</strong> the methods used to probe damage<br />

to the respiratory tract demonstrated that respiratory function data, blood<br />

gas measurements, and BALF analysis facilitate a meaningful interpretation<br />

<strong>of</strong> the effects observed and are important adjuncts to common inhalation<br />

toxicological studies on rats to describe quantitatively the diseased state <strong>of</strong><br />

the lung.<br />

The guinea-pig model is experimentally less demanding and therefore can<br />

suitably be used as a screening test for respiratory sensitization, as far as<br />

the limitations <strong>of</strong> this model are taken into account. Studies on guinea-pigs<br />

demonstrate that elicitation <strong>of</strong> respiratory hypersensitivity is<br />

challengeconcentration dependent and that the concentrations used should<br />

slightly exceed the threshold concentration for irritation to maximize the<br />

magnitude <strong>of</strong> the response. However, sensitization by inhalation may<br />

increase the susceptibility to irritant stimuli and thus confounds the<br />

selection <strong>of</strong> the most appropriate concentration for challenge. The<br />

combined approach <strong>of</strong> evaluating several breathing parameters, e.g.<br />

respiratory rate, flow- and volume-derived parameters, during both the<br />

hapten (free or conjugated) and the ACh challenge provides a promising<br />

method to distinguish specific and nonspecific hypersensitivity responses.<br />

Furthermore, it is critically important to assess the respiratory irritant<br />

potency <strong>of</strong> the compound under investigation. For potent irritant<br />

substances such as volatile isocyanates, challenge with the haptenprotein<br />

conjugate minimizes the likelihood to confound specific hypersensitivity<br />

responses with those evoked merely by irritation. Taking all imponderable<br />

factors into consideration, it appears that the guinea-pig intradermalinduction<br />

inhalation-challenge protocol is adequately susceptible to identify<br />

potent respiratory tract sensitizers. However, if the airway inflammation<br />

related features <strong>of</strong> asthma are the endpoints <strong>of</strong> primary interest other<br />

animal models appear to be more appropriate.<br />

References<br />

BARNES, P.J., BARANIUK, J.N. and BELVISI, M.G., 1991a, Neuropeptides in the<br />

respiratory tract (Part II). Am. Rev. Respir. Dis., 144, 1391–9.

J.PAULUHN 137<br />

BARNES, P.J., CHUNG, K.F., PAGE, C.P., 1991b, Pharmacology <strong>of</strong> asthma,<br />

Chapter 3, Inflammatory Mediators in Page, C.P. and Barnes, P.J. (Eds.), pp<br />

54–106. Handbook <strong>of</strong> Experimental Pharmacology, Berlin, Heidelberg, New<br />

York: Springer-Verlag.<br />

BOTHAM, P.A., HEXT, P.M., RATTRAY, N.J., WALSH, S.T. and WOOD-<br />

COCK, D.R., 1988, Sensitisation <strong>of</strong> guinea-pigs by inhalation exposure to lowmolecular-weight<br />

chemicals, Toxicol. Lett., 41, 159–73.<br />



and NIESSEN, H.J., 1993, Respiratory Allergy. ECETOC Monograph No. 19.<br />

BROOKS, S.M., WEISS, M.A. and BERNSTEIN, I.L., 1985, Reactive airway dysfunction<br />

syndrome (RADS). Persistent asthma syndrome after high level irritant<br />

exposure, Chest, 88, 376–84.<br />


A.J. and CHUNG, K.F., 1992, Bronchial hyperreactivity after inhalation <strong>of</strong><br />

trimellitic anhydride dust in guinea-pigs after intradermal sensitization to the<br />

free hapten, Am. Rev. Respir. Dis., 146, 1311–14.<br />

KAY, A.B., 1991, Asthma and inflammation, J. Allergy Clin. Immunol., 87, 893–<br />

910.<br />

KIPS, J.C., CUVELIER, C.A. and PAUWELS, R.A., 1992, Effect <strong>of</strong> acute and<br />

chronic antigen inhalation on airway morphology and responsiveness in<br />

actively sensitized rats, Am. Rev. Respir. Dis., 145, 1306–10.<br />

NIELSEN, G.D., 1991, Mechanisms <strong>of</strong> activation <strong>of</strong> the sensory irritant receptor by<br />

airborne chemicals, Crit. Rev. Toxicol, 21, 183–208.<br />

PAULUHN, J., 1994, Test methods for respiratory sensitization in use <strong>of</strong><br />

mechanistic information in risk assessment, EUROTOX Proceedings, Arch.<br />

Toxicol., suppl. 16, 77–86.<br />

PAULUHN, J. and EBEN, A., 1991, Validation <strong>of</strong> a non-invasive technique to<br />

assess immediate or delayed onset <strong>of</strong> airway hypersensitivity in guinea-pigs, J.<br />

Appl. Toxicol, 11, 423–31.<br />

PAULUHN, J. and MOHR, U., 1994, Assessment <strong>of</strong> respiratory hypersensitivity in<br />

guinea-pigs sensitized to diphenylmethane-4,4'-diisocyanate (MDI) and<br />

challenged with MDI, acetylcholine or MDI-albumin conjugate, <strong>Toxicology</strong> (in<br />

press).<br />


ALARIE, Y., 1993, Characteristic modifications <strong>of</strong> the breathing pattern <strong>of</strong><br />

mice to evaluate the effects <strong>of</strong> airborne chemicals on the respiratory tract, Arch.<br />

Toxicol, 67, 478–90.

10<br />

Mechanisms <strong>of</strong> Pulmonary Sensitization<br />


Zeneca Central <strong>Toxicology</strong> Laboratory, Macclesfield<br />

Introduction<br />

A wide range <strong>of</strong> chemicals is known to cause allergic contact dermatitis. It<br />

is apparent, however, that chemicals also have the potential to provoke other<br />

forms <strong>of</strong> allergy and <strong>of</strong> growing concern is pulmonary sensitization.<br />

Examples <strong>of</strong> chemicals identified as human respiratory allergens are listed<br />

in Table 10.1. Respiratory allergic hypersensitivity is characterized by<br />

pulmonary reactions which occur normally in only a proportion, and<br />

frequently in only a small proportion, <strong>of</strong> exposed individuals. In those who<br />

are sensitized, respiratory reactions can be provoked by atmospheric<br />

concentrations <strong>of</strong> the causative chemical allergen which were tolerated<br />

previously and which are without effect in the non-sensitized population<br />

(Newman Taylor, 1988). Almost invariably there is a latent period between<br />

the onset <strong>of</strong> exposure and the development <strong>of</strong> respiratory symptoms such<br />

as asthma and rhinitis.<br />

By definition, allergy, including sensitization <strong>of</strong> the respiratory tract,<br />

results from the stimulation <strong>of</strong> specific immune responses by the causative<br />

agent. Although it is assumed frequently that effective allergic sensitization<br />

<strong>of</strong> the respiratory tract results largely or wholly from inhalation exposure,<br />

this is not necessarily the case. Allergic reactions manifest in a particular<br />

organ commonly result from the local provocation by the inducing agent <strong>of</strong><br />

a systemically sensitized individual. There is no reason to suppose that the<br />

quality <strong>of</strong> immune response necessary for sensitization <strong>of</strong> the respiratory<br />

tract may not result from exposure to the chemical allergen at a different<br />

site. Consistent with this is evidence that occupational respiratory allergy<br />

may be caused by dermal contact with the chemical (Karol, 1986; Nemery<br />

and Lenaerts, 1993). Furthermore, it has been reported that respiratory<br />

rate changes can be provoked by inhalation exposure <strong>of</strong> guinea pigs<br />

sensitized previously by either topical or subcutaneous treatment with the<br />

same chemical (Karol et al., 1981; Rattray et al., 1994). Despite the fact<br />

that, in practice, pulmonary sensitization may not be caused exclusively by

Table 10.1 Chemicals identified as human respiratory allergens<br />

I.KIMBER 139<br />

inhalation <strong>of</strong> the chemical allergen, it is likely that this is an important<br />

route <strong>of</strong> exposure in the occupational setting.<br />

It is well established that respiratory sensitization caused by protein<br />

aeroallergens is effected by IgE antibody. This class <strong>of</strong> antibody in man is<br />

homocytotropic and is able to associate, via specific membrane receptors,<br />

with mast cells, including mast cells in the respiratory tract. Following<br />

subsequent exposure <strong>of</strong> the sensitized individual to the same allergen, mast<br />

cell-bound IgE is cross-linked and this, in turn, results in mast cell<br />

degranulation and the release <strong>of</strong> both preformed and newly-synthesized<br />

mediators which provoke acute inflammatory reactions. In the case <strong>of</strong><br />

sensitization <strong>of</strong> the respiratory tract caused by chemicals, however, an<br />

invariable association with the presence <strong>of</strong> specific IgE antibody has failed<br />

to emerge. Although IgE antibody specific for all recognized chemical<br />

respiratory allergens has been demonstrated, it is not uncommonly the case<br />

that individuals displaying symptoms <strong>of</strong> pulmonary hypersensitivity have<br />

been reported to lack demonstrable IgE. This may suggest that<br />

immunological processes independent <strong>of</strong> IgE antibody may play a decisive<br />

role in the induction <strong>of</strong> respiratory sensitization. An alternative explanation<br />

is that inappropriate or insensitive techniques have been employed for<br />

serological analysis and that IgE antibody may be associated more<br />

commonly than suspected previously with chemical respiratory allergy. In<br />

this context it is relevant that it has been found that positive skin prick<br />

tests can be provoked in patients sensitized to acid anhydrides who, on the<br />

basis <strong>of</strong> radioallergosorbent tests (RAST), were found to lack measurable<br />

levels <strong>of</strong> serum IgE antibody (Drexler et al., 1993). Despite the absence <strong>of</strong><br />

formal confirmation that there exists a universal causal relationship<br />

between specific IgE and pulmonary hypersensitivity induced by chemicals,


it remains likely that this class <strong>of</strong> antibody is responsible, in at least the<br />

majority <strong>of</strong> cases, for the acute onset symptoms associated with respiratory<br />

allergy (Karol et al., 1994).<br />

The induction and regulation <strong>of</strong> IgE responses<br />

IgE antibody responses are subject to a variety <strong>of</strong> immunoregulatory<br />

control mechanisms. Chief among these are the stimulatory and inhibitory<br />

actions <strong>of</strong> cytokines which serve to influence the induction and duration <strong>of</strong><br />

IgE responses. It has been found in mice that interleukin 4 (IL-4) is<br />

necessary for the initiation and maintenance <strong>of</strong> IgE antibody production<br />

(Finkelman et al., 1988b). The essential role for this cytokine in IgE<br />

responses has been emphasized further by studies <strong>of</strong> mice homozygous for<br />

a mutation that inactivates the gene for IL-4. These animals lack detectable<br />

serum IgE and fail to mount IgE responses (Kuhn et al., 1991). Importantly,<br />

in mice which produce constitutively high levels <strong>of</strong> IL-4, significantly<br />

elevated concentrations <strong>of</strong> serum IgE are evident (Burstein et al., 1991). A<br />

balance to the promotional influence <strong>of</strong> IL-4 is provided by interferon<br />

(IFN- ), a cytokine which exerts an inhibitory affect on IgE responses<br />

(Finkelman et al., 1988a). The reciprocal antagonistic activity <strong>of</strong> these<br />

cytokines is not restricted to the mouse, IL-4 and IFN- have been found to<br />

regulate human IgE production (Del Prete et al., 1988; Pene et al., 1988).<br />

The cytokines which influence the integrity <strong>of</strong> IgE responses are the<br />

products <strong>of</strong> discrete subpopulations <strong>of</strong> T helper (Th) cells, lymphocytes<br />

characterized by possession <strong>of</strong> the CD4 membrane determinant. It has been<br />

found in both mouse and man that there exists a functional heterogeneity<br />

among Th cells. Two major populations, designated Th 1 and Th 2, have<br />

been described (Mosmann and C<strong>of</strong>fman, 1989; Romagnani, 1991). It is<br />

believed currently that these subsets represent the most differentiated forms<br />

<strong>of</strong> Th cells and develop from less mature precursors as the immune<br />

response evolves (Mosmann et al., 1991). The major functional distinction<br />

between Th 1 and Th 2 cells resides in the spectrum <strong>of</strong> cytokines which they<br />

elaborate (Mosmann and C<strong>of</strong>fman, 1989). The cytokine products <strong>of</strong><br />

murine Th 1 and Th 2 cells are displayed in Table 10.2.<br />

It has been reported previously that chemicals known to cause<br />

respiratory hypersensitivity in man induce in mice immune responses<br />

characteristic <strong>of</strong> Th 2 cell activation, stimulate the production <strong>of</strong> specific IgE<br />

antibody and cause an increase in the serum concentration <strong>of</strong> IgE.<br />

Conversely, chemical allergens considered not to cause respiratory<br />

sensitivity, but which are nevertheless able to induce skin sensitization,<br />

elicit instead Th 1-type responses. In the latter case, immune responses are<br />

characterized by comparatively high levels <strong>of</strong> IgG2a antibody (an isotype<br />

known to be upregulated by IFN- ) and the absence <strong>of</strong> specific IgE<br />

(Dearman and Kimber, 1991, 1992; Dearman et al., 1991, 1992a,c,d,

Table 10.2 The cytokine products <strong>of</strong> murine Th 1 and Th 2 cells<br />

From: Mosmann and C<strong>of</strong>fman (1989).<br />

I.KIMBER 141<br />

1994). The implication is that certain chemicals favour the development <strong>of</strong><br />

Th 2 cells which will then synthesize and secrete IL-4 and thereby encourage<br />

IgE antibody responses and mast cell sensitization. The converse is that<br />

other classes <strong>of</strong> chemical allergen preferentially stimulate Th1 cells and IFNproduction.<br />

Such conditions will be nonpermissive for IgE antibody<br />

production and cell-mediated immune responses, including contact<br />

sensitization, will be favoured instead. A selective stimulation by different<br />

classes <strong>of</strong> chemical sensitizers <strong>of</strong> divergent Th responses may provide an<br />

explanation at the cellular level for the observation that chemicals vary<br />

with respect to the nature <strong>of</strong> allergic reactions that they will elicit<br />

preferentially in man. The stimulation by chemical allergens <strong>of</strong><br />

differentiated Th cell responses may have implications for allergic disease<br />

other than the regulation <strong>of</strong> IgE antibody. It is known for instance that<br />

IL-3, IL-4 and IL-10, all <strong>of</strong> which are products <strong>of</strong> murine Th 2 cells<br />

(Table 10.2), act as mast cell growth factors or c<strong>of</strong>actors (Smith and<br />

Rennick, 1986; Thompson-Snipes et al., 1991). Moreover, IL-5 is a growth<br />

and differentiation factor for eosinophils (Yokota et al., 1987) and serves<br />

to regulate the accumulation <strong>of</strong> these cells at the site <strong>of</strong> allergeninduced<br />

hypersensitivity reactions in the respiratory tract (Gulbenkian et al., 1992).<br />

It has been found recently that the cytokines IL-3 and IL-4 also enhance the<br />

secretory activity <strong>of</strong> mast cells following activation (Coleman et al., 1993).<br />

Antagonistic and inhibitory influences <strong>of</strong> Th cell products may also affect<br />

the elicitation <strong>of</strong> allergic reactions. It has been found that IFN- not only<br />

suppresses the secretory function <strong>of</strong> mast cells (Holliday et al., 1994), but<br />

also antagonizes the antigen-induced infiltration <strong>of</strong> eosinophils into the<br />

respiratory tract <strong>of</strong> sensitized mice (Iwamoto et al., 1993). Contact allergic<br />

reactions may in theory be regulated by Th 2 cytokines. It has been shown<br />

that IL-4 and IL-10 act in concert to inhibit Th 1 cell function and to


depress cell-mediated immunity (Powrie et al., 1993) and that Il–4 is able<br />

to reduce significantly the severity <strong>of</strong> contact allergic reactions in mice<br />

(Gautam et al., 1992).<br />

Taken together the available data suggest that the selective stimulation<br />

<strong>of</strong> Th cell responses and the consequent balance created between Th 1- and<br />

Th 2-derived cytokines will have an important impact on both the induction<br />

and elicitation stages <strong>of</strong> allergy. It is perhaps not surprising, therefore, that<br />

there is increasing evidence for selective Th responses in human allergic<br />

disease. Clones <strong>of</strong> T lymphocytes specific for aeroallergens such as house<br />

dust mite and grass pollen, which cause IgE-mediated respiratory allergic<br />

reactions in susceptible individuals, have been shown to elaborate Th 2<br />

cytokines, but not IFN- (Parronchi et al., 1991). A predominance <strong>of</strong> the<br />

Th 2-type cells has been found at sites <strong>of</strong> skin reactions in atopic individuals<br />

(Kay et al., 1991) and increased numbers <strong>of</strong> IL-4 + T lymphocytes have been<br />

identified in the nasal mucosa in allergen-induced rhinitis (Ying et al., 1994).<br />

By contrast, human immune responses to nickel, a common cause <strong>of</strong><br />

allergic contact dermatitis, are characterized by the selective activation <strong>of</strong><br />

Th 1-type cells. Allergen-specific T lymphocyte clones isolated from the<br />

peripheral blood <strong>of</strong> patients sensitized to nickel have been found to secrete<br />

only low or undetectable amounts <strong>of</strong> IL-4 and IL-5, but high levels <strong>of</strong> IFN-<br />

(Kapsenberg et al., 1991).<br />

Although the relative contribution <strong>of</strong> Th 1 and Th 2 cells during immune<br />

responses, and in particular the relative availability <strong>of</strong> IL-4 and IFN- , is<br />

likely to play a predominant role in the regulation <strong>of</strong> IgE antibody, other<br />

factors may be relevant. Not least, the priming <strong>of</strong> Th 1 cells for the<br />

production <strong>of</strong> IFN- may in turn be dependent upon the action <strong>of</strong> another<br />

cytokine, interleukin 12 (IL-12) (Manetti et al., 1994; Morris et al., 1994;<br />

Schmitt et al., 1994). It has been demonstrated also that CD8 + T<br />

lymphocytes exert an important immunoregulatory influence on IgE<br />

responses (Kemeny et al., 1994; Renz et al., 1994), possibly via downregulation<br />

<strong>of</strong> CD4 + Th 2 cell development (Noble et al., 1993).<br />

It is clear that conditions outwith the immune system also influence the<br />

magnitude <strong>of</strong> IgE responses. Certainly genetic predisposition is an<br />

important, although poorly understood factor. In addition, there have been<br />

suggestions that cigarette smoking and exposure to certain environmental<br />

pollutants may result in increased IgE levels and may also serve to<br />

aggravate asthma (Zetterstrom et al., 1981; Muranka et al., 1986;<br />

Wardlaw, 1993).<br />

Cell-mediated immune responses in chemical respiratory<br />

allergy<br />

The elicitation <strong>of</strong> chemical respiratory hypersensitivity may be associated<br />

with both immediate-onset and late phase reactions. While IgE antibody

and local degranulation <strong>of</strong> mast cells may be necessary for acute<br />

symptoms, late asthmatic responses appearing some hours following<br />

exposure are characterized by an infiltration <strong>of</strong> mononuclear cells and<br />

increased numbers <strong>of</strong> leucocytes in bronchoalveolar lavage fluid. Chronic<br />

inflammation is an important component <strong>of</strong> asthma and, in addition to<br />

mononuclear cell accumulation, is characterized by mucus production, the<br />

destruction and sloughing <strong>of</strong> airway epithelial cells and subepithelial<br />

fibrosis secondary to collagen deposition. Eosinophils, acting together with<br />

infiltrating T lymphocytes, play a pivotal role in chronic bronchial<br />

inflammation (Corrigan and Kay, 1992). It is apparent also that the<br />

generation <strong>of</strong> eosinophilia in the respiratory tract is influenced markedly by<br />

Th cell products. As described previously, IL-5 effects the accumulation <strong>of</strong><br />

eosinophils at the site <strong>of</strong> hypersensitivity reactions in respiratory tissues,<br />

while IFN- , secondary to an inhibition <strong>of</strong> CD4 + cell infiltration,<br />

antagonizes this process (Gulbenkian et al., 1992; Iwamoto et al., 1993). It<br />

may prove that the cell-mediated immune processes relevant to the<br />

development <strong>of</strong> respiratory hypersensitivity and asthma are also a function<br />

<strong>of</strong> Th cell heterogeneity. Certainly the stimulation <strong>of</strong> Th 2 cell activation<br />

will have pr<strong>of</strong>ound effects on all stages <strong>of</strong> respiratory allergy. The<br />

infiltration <strong>of</strong> such cells into sites <strong>of</strong> encounter with inducing allergen, a<br />

process perhaps facilitated by vasodilation resulting from mast cell<br />

degranulation, will provide a local source <strong>of</strong> cytokines such as IL-4 and<br />

IL-5. Mast cell secretory activity will be potentiated by the former and<br />

eosinophil accumulation triggered by the latter. That Th 2 cells do in fact<br />

accumulate in the area <strong>of</strong> immediate-type hypersensitivity reactions is<br />

supported by the studies <strong>of</strong> Kay et al. (1991) who demonstrated that the<br />

cells infiltrating lesional skin at the sites <strong>of</strong> late phase cutaneous reactions<br />

in atopic patients produce IL-3, IL-4, IL-5 and GM-CSF, but not IFN- .<br />

Practical applications<br />

I.KIMBER 143<br />

In the course <strong>of</strong> investigations designed to examine the characteristics <strong>of</strong><br />

immune responses induced in mice by chemical sensitizers it was found<br />

that only those materials known to cause respiratory hypersensitivity in<br />

man provoked in mice a substantial increase in the serum concentration <strong>of</strong><br />

IgE; a phenomenon thought to reflect the selective stimulation <strong>of</strong> Th 2 celltype<br />

responses by this class <strong>of</strong> allergen. It was observed also that contact<br />

allergens known or suspected not to cause occupational respiratory<br />

hypersensitivity failed to result in similar changes in serum IgE levels<br />

(Dearman and Kimber, 1991, 1992; Dearman et al., 1992a,d). The<br />

differential ability <strong>of</strong> chemical respiratory and contact allergens to<br />

stimulate changes in the concentration <strong>of</strong> serum IgE in mice forms the basis<br />

<strong>of</strong> a novel approach to the identification <strong>of</strong> chemicals which have the<br />

potential to cause sensitization <strong>of</strong> the respiratory tract. This method, the


mouse IgE test (Dearman et al., 1992b, Kimber and Dearman, 1993) is<br />

being evaluated currently in the context <strong>of</strong> internal and inter-laboratory<br />

validation studies.<br />

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DEARMAN, R.J., BASKETTER, D.A., COLEMAN, J.W. and KIMBER, I., 1992a,<br />

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DEARMAN, R.J., BASKETTER, D.A. and KIMBER, I., 1992b, Variable effects <strong>of</strong><br />

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DEARMAN, R.J., MITCHELL, J.A., BASKETTER, D.A. and KIMBER, I., 1992c,<br />

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to cause immediate and delayed dermal hypersensitivity reactions in mice,<br />

International Archives <strong>of</strong> Allergy and Immunology, 97, 315–21.

I.KIMBER 145<br />

DEARMAN, R.J., SPENCE, L.M. and KIMBER, I., 1992d, Characterization <strong>of</strong><br />

murine immune responses to allergenic diisocyanates, <strong>Toxicology</strong> and Applied<br />

Pharmacology, 112, 190–7.<br />

DEARMAN, R.J., RAMDIN, L.S.P., BASKETTER, D.A. and KIMBER, I. 1994,<br />

Inducible interleukin-4-secreting cells provoked in mice during chemical<br />

sensitization, Immunology, 81, 551–7.<br />


MACCHIA, D., RICI, M., ANSARI, A.A. and ROMAGNANI, S., 1988, IL-4<br />

is an essential factor for the IgE synthesis induced in vitro by human T cell<br />

clones and their supernatants, Journal <strong>of</strong> Immunology, 140, 4193–8.<br />



1987, Clinical and immunological investigations <strong>of</strong> respiratory disease in<br />

workers using reactive dyes, British Journal <strong>of</strong> lndustrial Medicine, 44, 534–41.<br />


1993, Skin prick tests with solutions <strong>of</strong> acid anhydrides in acetone,<br />

International Archives <strong>of</strong> Allergy and Immunology, 100, 251–5.<br />


1988a, IFN- regulates the isotypes <strong>of</strong> Ig secreted during in vivo humoral<br />

immune responses, Journal <strong>of</strong> Immunology, 140, 1022–7.<br />


TUNG, A.S., SAMPLE, J.G. and PAUL, W.E., 1988b, IL-4 is required to<br />

generate and sustain in vivo IgE responses, Journal <strong>of</strong> Immunology, 141, 2335–<br />

41.<br />

GAUTAM, S.C., CHIKKALA, N.F. and HAMILTON, T.A., 1992,<br />

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trinitrochlorobenzene, Journal <strong>of</strong> Immunology, 148, 1411–15.<br />



WATNIK, A.S., 1992, Interleukin-5 modulates eosinophil accumulation in<br />

allergic guinea pig lung, American Review <strong>of</strong> Respiratory Diseases, 146, 263–9.<br />


COLEMAN, J.W., 1994. Interactions <strong>of</strong> IFN- with IL-3 and IL-4 in the<br />

regulation <strong>of</strong> serotonin and arachidonate release from peritoneal mast cells,<br />

Immunology, 82, 70–4.<br />


LAW, J.S. and NEWMAN TAYLOR, A.J., 1983, Tetrachlorophthalic<br />

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Clinical Immunology, 7l, 5–11.<br />

IWAMOTO, I., NAKAJIMA, H., ENDO, H. and YOSHIDA, S., 1993, Interferon<br />

regulates antigen-induced eosinophil recruitment into the mouse airways by<br />

inhibiting infiltration <strong>of</strong> CD4 + T cells, Journal <strong>of</strong> Experimental Medicine, 177,<br />

573–6.<br />

KAPSENBERG, M.L., WIERENGA, E.A., Bos, J.D. and JANSEN, H.M., 1991,<br />

Functional subsets <strong>of</strong> allergen-reactive human CD4 + T cells, Immunology<br />

Today, 12, 392–5.<br />

KAROL, M.H., 1986, Respiratory effects <strong>of</strong> inhaled isocyanates, CRC Critical<br />

Reviews in <strong>Toxicology</strong>, 16, 349–79.


KAROL, M.H., HAUTH, B.A., RILEY, E.J. and MAGRENI, C.M., 1981, Dermal<br />

contact with toluene diisocyanate (TDI) produces respiratory tract<br />

hypersensitivity in guinea pigs, <strong>Toxicology</strong> and Applied Pharmacology, 58,<br />

221–30.<br />


P., SAETTA, M. and MAPP, C.E., 1994, Predictive value <strong>of</strong> airways hyperresponsiveness<br />

and circulating IgE for identifying types <strong>of</strong> responses to toluene<br />

diisocyanate inhalation challenge, American Journal <strong>of</strong> Respiratory and<br />

Critical Care Medicine, 149, 611–15.<br />

KAY, A.B., YING, S., VARNEY, V., GAGA, M., DURHAM, S.R., MOQBEL, R.,<br />

WARDLAW, A.J. and HAMID, Q., 1991, Messenger RNA expression <strong>of</strong> the<br />

cytokine gene cluster, interleukin 3 (IL-3), IL-4, IL-5 and granulocyte/<br />

macrophage colony stimulating factor, in allergen-induced last phase reactions<br />

in atopic subjects, Journal <strong>of</strong> Experimental Medicine, 173, 775–8.<br />

KEMENY, D.M., NOBLE, A., HOLMES, B.J. and DIAZ-SANCHEZ, D., 1994,<br />

Immune regulation: a new role for the CD8 + T cell, Immunology Today, 15,<br />

107–10.<br />

KIMBER, I. and DEARMAN, R.J., 1993, Approaches to the identification and<br />

classification <strong>of</strong> chemical allergens in mice, Journal <strong>of</strong> Pharmacological and<br />

Toxicological Methods, 29, 11–16.<br />

KUHN, R., RAJEWSKY, K. and MULLER, W., 1991, Generation and analysis <strong>of</strong><br />

interleukin-4 deficient mice, Science, 254, 707–10.<br />

MACCIA, C.A. BERNSTEIN, I.L., EMMETT, E.A. and BROOKS, S.M. 1976, In<br />

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TRI NCHIERI, G., 1994, Interleukin 12 induces stable priming for interferon<br />

(IFN- ) production during differentiation <strong>of</strong> human T helper (Th) cells and<br />

transient IFN- production in established Th 2 cell clones, Journal <strong>of</strong><br />

Experimental Medicine, 179, 1273–83.<br />


BURROUGHS, H.E. and BERNSTEIN, I.L., 1985, Detection <strong>of</strong> IgE-mediated<br />

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Journal <strong>of</strong> Allergy and Clinical Immunology, 15, 663–72.<br />


B.R., GATELY, M.K. and FINKELMAN, F.D., 1994, Effects <strong>of</strong> IL-12 on in<br />

vivo cytokine gene expression and Ig isotype selection, Journal <strong>of</strong><br />

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MOSMAN, T.R. and COFFMAN, R.L., 1989, Heterogeneity <strong>of</strong> cytokine secretion<br />

patterns and functions <strong>of</strong> helper T cells, Advances in Immunology, 46, 111–<br />

47.<br />


A., FONG, T.A.T., BOND, M.W., MOORE, K.W.M., SHER, A. and<br />

FIORENTINO, D.F., 1991, Diversity <strong>of</strong> cytokine synthesis and function <strong>of</strong><br />

mouse CD4 + T cells, Immunological Reviews, 123, 209–29.<br />


IKEMURI, R. and TOKIWA, H., 1986, Adjuvant activity <strong>of</strong> diesel-exhaust

I.KIMBER 147<br />

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Clinical Immunology, 77, 616–23.<br />

MURDOCH, R.D., PEPYS, J. and HUGHES, E.G., 1986. IgE antibody responses<br />

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<strong>Industrial</strong> Medicine, 43, 37–43.<br />

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diisocyanate in coal mines, Lancet, 341, 318.<br />

NEWMAN TAYLOR, A.J., 1988, Occupational asthma, Postgraduate Medical<br />

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M., 1993, Elimination <strong>of</strong> IgE regulatory rat CD8 + T cells in vivo increases the<br />

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BANCHEREAU, J., SPITS, H. and DE VRIES, J.E., 1988, IgE production by<br />

normal human B cells induced by alloreactive T cell clones is mediated by 11–4<br />

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P., LUH, H., HOY, P., PENE, J., BRIERE, F., SPITS, H., BANCHEREAU, J.,<br />

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11<br />

Occupational Asthma Induced by Chemical<br />

Agents<br />


Wythenshawe Hospital Manchester<br />

Introduction<br />

Occupational asthma may be defined as variable airways narrowing<br />

causally related to exposure in the working environment to airborne dust,<br />

gases, vapours or fumes (Newman Taylor, 1980). This definition includes,<br />

therefore, both immunological and nonimmunological causes <strong>of</strong> asthma in<br />

the workplace. Immunological causes <strong>of</strong> asthma in general demonstrate a<br />

latent period between exposure and the development <strong>of</strong> symptoms. Once<br />

sensitisation has occurred airway responses may be seen at very low levels<br />

<strong>of</strong> exposure. Both high and low molecular weight agents may cause<br />

sensitisation. Irritantinduced occupational asthma characteristically follows<br />

within 24 h <strong>of</strong> a usually, single, high level exposure to an irritant substance<br />

and has been named reactive airways dysfunction syndrome (Brooks et al.,<br />

1985).<br />

The number <strong>of</strong> chemical agents causing occupational asthma is<br />

extensive. As new, highly reactive, chemicals are developed these numbers<br />

are likely to grow. Low molecular weight chemicals may act as haptens,<br />

reacting with body protein to form a complete antigen to which specific<br />

antibodies are formed.<br />

The incidence <strong>of</strong> occupational asthma in most countries is not known<br />

with any great accuracy, there are considerable variations in reporting<br />

systems between countries. Since many individuals with occupational<br />

asthma change jobs without a specific diagnosis being established, the<br />

published figures <strong>of</strong> incidence will be significant underestimates <strong>of</strong> the true<br />

incidences. In Japan (Kobayashi et al., 1973), the prevalence <strong>of</strong><br />

occupational asthma amongst adult male asthmatics is said to be about<br />

15%. In the UK a new reporting system has recently been established—<br />

Surveillance <strong>of</strong> Work-related and Occupational Respiratory Disease Project<br />

(SWORD). Newly diagnosed cases <strong>of</strong> workrelated respiratory disease are<br />

reported monthly by consultant chest and occupational physicians.<br />

Between 1989 and 1991, 631 cases <strong>of</strong> chemically induced occupational<br />

asthma were reported, <strong>of</strong> these 53% were associated with exposure to


isocyanates. A similar system (SHIELD) has been established in the UK in<br />

the West Midlands (Gannon et al., 1993). They reported an incidence <strong>of</strong> 43<br />

new cases per million workers per year. Specific occupational incidences<br />

varied from 1833 per million paint sprayers to 8 per million clerks. Again<br />

more than half the cases <strong>of</strong> asthma were attributed to isocyanates.<br />

The initial diagnosis <strong>of</strong> occupational asthma is based on a workers<br />

history <strong>of</strong> respiratory symptoms improving on days away from work and<br />

when on holiday. At the onset <strong>of</strong> occupational asthma this pattern is<br />

usually present but continued exposure to the allergen leads to increasing<br />

airway reactivity. Their symptoms may then persist over weekends and be<br />

triggered by nonspecific factors outside the workplace such as exhaust<br />

fumes, aerosol sprays and perfumes. The difficulties <strong>of</strong> relying on the<br />

history alone in the diagnosis <strong>of</strong> occupational asthma has been well<br />

documented (Malo et al., 1991). A series <strong>of</strong> 162 hospital referrals to two<br />

expert physicians were initially categorised, on the basis <strong>of</strong> their histories,<br />

into highly probable, probable, uncertain, unlikely or absent occupational<br />

asthma. The diagnosis was then established by bronchial provocation<br />

testing and or serial measurements <strong>of</strong> lung function. The predictive value <strong>of</strong><br />

a physician’s assessment <strong>of</strong> occupational asthma being highly probable or<br />

probable was only 63%. This improved to 83% in the groups in whom<br />

occupational asthma was assessed as being unlikely or absent.<br />

The early identification <strong>of</strong> work-related symptoms and their subsequent<br />

investigation in the workplace is important. Only rarely, when very acute<br />

episodes <strong>of</strong> workplace asthma are described, should lung function<br />

measurements at work be avoided. While pre- and post-shift measurements<br />

<strong>of</strong> lung function may identify a work-related effect, late asthmatic<br />

responses occurring in the evening after leaving work are frequently seen in<br />

chemically induced forms <strong>of</strong> occupational asthma. Serial measurements <strong>of</strong><br />

lung function made every 2 h from waking to sleeping, both on working<br />

days and on days away from work, using a peak flow meter, will identify<br />

these late responders. The sensitivity <strong>of</strong> this type <strong>of</strong> investigation in<br />

establishing a diagnosis <strong>of</strong> occupational asthma is about 80% (Burge, 1982),<br />

this falls to 46% once the worker is started on specific treatment for his<br />

asthma, again emphasising the importance <strong>of</strong> early identification and<br />

investigation <strong>of</strong> work-related respiratory symptoms.<br />

Currently more than 140 low molecular weight chemicals have been<br />

reported to induce occupational asthma (Butcher and Salvaggio, 1986).<br />

The majority <strong>of</strong> these chemicals induce asthma by mechanisms which have<br />

yet to be identified. In a minority <strong>of</strong> instances specific IgE antibodies to the<br />

implicated chemical have been identified.<br />

Bronchial challenge tests with chemicals which are non-IgE dependent<br />

usually induce either an isolated late asthmatic response or a biphasic or<br />

dual asthmatic response. The IgE dependent responses induce immediate or<br />

dual asthmatic responses.

The most common chemical causes <strong>of</strong> occupational asthma include the<br />

iso- cyanates and the acid anhydrides. This chapter will examine these two<br />

groups in more detail.<br />

Isocyanates<br />

C.A.C.PICKERING 151<br />

The polyisocyanates and their oligomers are the most important cause <strong>of</strong><br />

chemically induced asthma. These organic compounds are synthesised by<br />

the reaction <strong>of</strong> amines with phosgene. There are a number <strong>of</strong> related<br />

compounds the most important <strong>of</strong> which are 2,4- and 2,6-toluene<br />

diisocyanate (TDI), methylene diphenyldiisocyanate (MDI), hexamethylene<br />

diisocyante (HDI), napthalene diisocyanate (NDI), isophorone diisocyanate<br />

(IPDI), and polyisocyanates derived from HDI and MDI.<br />

The incidence <strong>of</strong> occupational asthma due to diisocyanates varies widely.<br />

It is influenced by the type <strong>of</strong> compound and its vapour pressure. TDI and<br />

HDI are highly volatile at room temperature, whereas MDI has to be<br />

heated to above 60°C to volatilise. It is thought that approximately 5% <strong>of</strong><br />

an exposed working population will develop occupational asthma after<br />

exposure to TDI (Diem et al., 1982). Because <strong>of</strong> the known respiratory<br />

problems associated with exposure to isocyanates with high vapour<br />

pressure properties, new isocyanate compounds with low vapour pressure<br />

properties have been developed particularly for use in the paint spraying<br />

industry. Recent studies however continue to demonstrate significant levels<br />

<strong>of</strong> occupational asthma despite the use <strong>of</strong> recommended respiratory<br />

protection (Seguin et al., 1987, Welinder et al., 1988). Bronchial<br />

provocation studies with HDI- and MDI-derived polyisocyanates have<br />

confirmed their ability to cause occupational asthma. Airborne iso-cyanate<br />

prepolymers appear to be able to induce asthma to the same or greater<br />

frequency as isocyanate monomers.<br />

High exposures to isocyanate vapours, such as occur in a major<br />

industrial spillage, cause acute rhinitis, lacrymation, cough and wheezing<br />

leading to subsequent sensitisation. In some individuals this type <strong>of</strong><br />

exposure induces persistent asthma—reactive airways dysfunction<br />

syndrome (RADS). Respiratory sensitisation may occur at very low levels<br />

<strong>of</strong> exposure. Pepys et al. (1972) described a boat builder who became<br />

sensitised to TDI at exposure levels <strong>of</strong> between 0.00173 and 0.0018 ppm.<br />

Similarly White et al. (1980) reported respiratory symptoms and the<br />

development <strong>of</strong> IgE antibodies to TDI, in machinists manufacturing carseat<br />

covers exposed to levels <strong>of</strong> TDI <strong>of</strong> between 0.0003 and 0.003 ppm. It is<br />

more usual, in the author’s experience, for the sensitised individual to<br />

provide a history <strong>of</strong> short lived peak exposures to isocyanates which have<br />

clearly been above the current threshold limit value. These intermittent<br />

relatively high level exposures may be important in the sensitisation


process. Once sensitised, a worker may have his symptoms initiated by very<br />

low exposure levels <strong>of</strong> isocyanates.<br />

Diisocyanate asthma is usually but not always associated with the<br />

presence <strong>of</strong> nonspecific bronchial hyperreactivity. The majority <strong>of</strong> workers<br />

who develop occupational asthma remain symptomatic requiring regular<br />

treatment permanently after cessation <strong>of</strong> exposure (Allard et al., 1989).<br />

The duration <strong>of</strong> exposure with symptoms before diagnosis has a major<br />

influence on recovery patterns. In a group <strong>of</strong> 43 isocyanate workers with<br />

occupational asthma, those who had fully recovered were exposed with<br />

symptoms for 1.6 years, those who had improved, 2.8 years and those who<br />

had not improved, 5.4 years (Pisati et al., 1993). The resolution or<br />

improvement in occupational asthma takes place over a 2 year period after<br />

cessation <strong>of</strong> exposure, symptoms still present at 2 years should be regarded<br />

as permanent. Most epidemiological studies have not identified any specific<br />

risk factors including atopic status, smoking or nonspecific bronchial<br />

hyperreactivity.<br />

The laboratory identification <strong>of</strong> specific antibodies to diisocyanates has<br />

proved <strong>of</strong> very limited value. Diisocyanate specific IgE is demonstrable in<br />

only 10–20% <strong>of</strong> sensitised individuals and have also been identified in<br />

individuals with no history <strong>of</strong> asthma (Butcher et al., 1983). Similarly<br />

specific IgG antibodies to diisocyanates have been described in workers<br />

both with and without evidence <strong>of</strong> disease.<br />

At the present time the recommended long-term exposure limit (8 h<br />

TWA reference period) for diisocyanates is 0.02 mg m −3 and the short-term<br />

exposure limit (10 min reference period) is 0.07 mg m −3 in the UK. There is<br />

discussion at the present time as to whether levels should be lower in order<br />

to prevent the development <strong>of</strong> diisocyanate asthma. However since most<br />

workers with diisocyanate airways disease describe exposures in excess <strong>of</strong><br />

the current recommended exposure levels the prevalence <strong>of</strong> occupational<br />

asthma in a workforce without such exposures is not known. There need to<br />

be improvements in hygiene control to prevent these peak exposures to<br />

isocyanates.<br />

Acid anhydrides<br />

The acid anhydrides are a group <strong>of</strong> low molecular weight chemicals used as<br />

curing agents in the production <strong>of</strong> epoxy and alkyd resins and in the<br />

production <strong>of</strong> plasticisers such as dioctyl phthalate. Acid anhydrides exert<br />

diverse effects on man both as sensitisers, irritants or both. The most<br />

frequently used anhydrides, all <strong>of</strong> which have been described causing<br />

occupational asthma, are phthalic anhydride (PA), trimellitic anhydride<br />

(TMA), tetrachlorophthalic anhydride (TCPA) and maleic anhydride<br />

(MA). In addition himic anhydride and pyromellitic dianhydride (PMDA)<br />

have been described as causing asthma.

The direct toxicity <strong>of</strong> anhydrides involves irritation <strong>of</strong> the mucus<br />

membranes and skin which may result in eye lesions, epistaxis, pulmonary<br />

congestion, haemoptysis and skin burns (Venables, 1989).<br />

Occupational asthma is most frequently reported due to PA, less<br />

commonly to TMA, TCPA and MA and finally there are single case reports<br />

<strong>of</strong> asthma due to HA, HHPA and PMDA (Venables, 1989). A second type<br />

<strong>of</strong> response to acid anhydrides has also been described and is termed the<br />

‘late respiratory systemic syndrome’ (LRRS). This is characterised by the<br />

development <strong>of</strong> influenzal type symptoms, fever, generalised acheing and<br />

malaise, late in the working shift or in the evening after work. These<br />

symptoms may occur in isolation or in association with asthma. It is not<br />

clear whether this response is immunologically mediated or a nonspecific<br />

response to high levels <strong>of</strong> anhydride exposure. Lastly, exposure to TMA,<br />

probably at high exposure levels, has been described as causing severe<br />

pulmonary haemorrhage requiring both blood transfusion and mechanical<br />

ventilation (Rivera et al., 1989).<br />

Serum IgE and IgG antibodies to acid anhydrides have been identified.<br />

IgE antibodies appear to be more specifically associated with occupational<br />

asthma. Howe et al. (1983) reported seven cases <strong>of</strong> TCPA asthma all <strong>of</strong><br />

whom had IgE antibody to TCPA, compared with 8% <strong>of</strong> 300 exposed<br />

workers without TCPA asthma; 29% <strong>of</strong> this exposed nonasthmatic<br />

population had IgG antibodies to TCPA.<br />

The exposure levels <strong>of</strong> acid anhydrides that initiate sensitisation are<br />

poorly understood. TMA at levels <strong>of</strong> 1.7–4.7 mg m −3 (Zeiss et al., 1977)<br />

and 0.007–2.1 mg m −3 (Bernstein et al., 1983; McGrath et al., 1984) have<br />

been described causing occupational asthma. PA at 0.03–15 mg m −3 has<br />

also been reported as causing asthma (Wernfors et al., 1986). As in other<br />

forms <strong>of</strong> occupational asthma, the early identification <strong>of</strong> cases <strong>of</strong> acid<br />

anhydride induced asthma and their removal from exposure is <strong>of</strong> prime<br />

importance.<br />

Reactive airways dysfunction syndrome<br />

C.A.C.PICKERING 153<br />

Reactive airways dysfunction syndrome (RADS) or irritant-induced asthma<br />

was first described in 1985 (Brooks et al., 1985). The criteria used in<br />

diagnosis include a high level exposure to an irritant fume, vapour or smoke,<br />

the development <strong>of</strong> respiratory symptoms within minutes or hours <strong>of</strong><br />

exposure, in an individual with no previous history <strong>of</strong> respiratory symptoms,<br />

with persistence <strong>of</strong> symptoms and physiological abnormalities for more<br />

than 1 year. A variety <strong>of</strong> different chemical exposures have been described<br />

inducing this syndrome including: chlorine (Moore and Sherman, 1991),<br />

glacial acetic acid (Kern, 1991), hydrochloric acid (Promisl<strong>of</strong>f et al., 1990)<br />

and miscellaneous chemical exposures (Brooks et al., 1985). A comparison<br />

between cases <strong>of</strong> occupational asthma and RADS (Gautrin et al., 1994)


suggest that cases with RADS are left with less airway reversibility than<br />

occupational asthmatics. This would be consistent with the pathological<br />

findings (Boutet et al., 1993) in RADS, with more severe basement<br />

membrane thickening and bronchial wall fibrosis than is present in<br />

occupational asthma.<br />

The development <strong>of</strong> occupational asthma in any individual has<br />

potentially serious consequences both in terms <strong>of</strong> persisting disability,<br />

possible unemployment and loss <strong>of</strong> income (Gannon et al., 1993). It is<br />

incumbent on management to ensure safe working conditions with<br />

adequate control and regular monitoring <strong>of</strong> atmospheric levels <strong>of</strong> chemical<br />

agents.<br />

References<br />

ALLARD, C., CARTIER, A., GHEZZO, H. and MALO, J-L., 1989, Occupational<br />

asthma due to various agents. Absence <strong>of</strong> clinical and functional improvement<br />

at an interval <strong>of</strong> four or more years after cessation <strong>of</strong> exposure, Chest, 96,<br />

1046–9.<br />


and PATTERSON, R., 1983, The relationship <strong>of</strong> airborne trimellitic<br />

anhydrideinduced symptoms and immune responses, J. Allergy Clin.<br />

Immunol., 72, 709–13.<br />


C., MILOT, J. and LAVIOLETTE, M., 1993, Morphological evidence <strong>of</strong><br />

modified contractile properties <strong>of</strong> airways in occupational asthma and reactive<br />

airways dys-function syndrome, Am. Rev. Respir. Dis., 147, A113.<br />

BROOKS, S.M., WEISS, M.A. and BERNSTEIN, I.L., 1985, Reactive airways<br />

dysfunction syndrome. Case reports <strong>of</strong> persistent airways hyperreactivity<br />

following high-level irritant exposures, J. Occup. Med., 27, 473–6.<br />

BURGE, P.S., 1982, Single and serial measurements <strong>of</strong> lung function in the<br />

diagnosis <strong>of</strong> occupational asthma, Eur. J. Resp. Dis., 63 (suppl. 123), 47–9.<br />

BUTCHER, B.T. and SALVAGGIO, J.E., 1986, Continuing medical education—<br />

occupational asthma, J. Allergy Clin. Immunol. 78, 547–9.<br />

BUTCHER, B.T., O’NEIL, C.E., REED, M.A. and SALVAGGIO, J.E., 1983, Radioallergosorbent<br />

testing with p-tolyl monoisocyanate in toluene diisocyanate<br />

workers, Clin. Allergy., 13, 31–4.<br />



1982, Five year longitudinal study <strong>of</strong> workers employed in a new toluene<br />

diisocyanate manufacturing plant, Am. Rev. Respir. Dis., 126, 420–8.<br />

GANNON, P.F.G. and BURGE, P.S., 1993, The SHIELD scheme in the West<br />

Midlands region, United Kingdom, Brit. J. Ind. Med., 50, 791–6.<br />

GANNON, P.F.G., WEIR, D.C., ROBERTSON, A.S. and BURGE, P.S., 1993,<br />

Health, employment, and flnancial outcomes in workers with occupational<br />

asthma, Brit. J. Ind. Med., 50, 491–6.

C.A.C.PICKERING 155<br />


L’ARCHEVÊQUE, J., LAVIOLETTE, M., CÔTÉ, J. and MALO, J.-L., 1994,<br />

Is reactive airways dysfunction syndrome a variant <strong>of</strong> occupational asthma? J.<br />

Allergy Clin. Immunol, 93, 12–22.<br />


LAW, J.S. and NEWMAN TAYLOR, A.J., 1983, Tetrachlorophthalic<br />

anyhdride asthma: evidence for specific IgE antibody, J. Allergy Clin.<br />

Immunol., 71, 5–11.<br />

KERN, D.G., 1991. Outbreak <strong>of</strong> the reactive airways dysfunction syndrome after a<br />

spill <strong>of</strong> glacial acetic acid, Am. Rev. Respir. Dis. 144, 1058–64.<br />


KISHIMOTO, T. and MIYAMOTO, T., 1973, Occupational asthma due to the<br />

inhalation <strong>of</strong> pharmaceutical dusts and other chemical agents with some<br />

reference to other occupational asthma in Japan, Proc. VIII Int. Congr.<br />

Allergology, Tokyo, October 1973, pp. 124–32. Amsterdam: Excerpta<br />

Medica.<br />


1991, Is the clinical history a satisfactory means for diagnosing occupational<br />

asthma? Am. Rev. Respir. Dis., 143 528–32.<br />

MCGRATH, K.G., ROACH, D., ZEISS, C.R. and PATTERSON, R., 1984,<br />

Four year evaluation <strong>of</strong> workers exposed to trimellitic anhydride: a brief<br />

report, J. Occup. Med., 26, 671–5.<br />

MOORE, B.B. and SHERMAN, M., 1991, Chronic reactive airway disease<br />

following acute chlorine gas exposure in an asymptomatic atopic patient,<br />

Chest, 100, 855–6.<br />

NEWMAN TAYLOR, A.J., 1980, Occupational asthma, Thorax, 35, 241–5.<br />

PEPYS, J., PICKERING, C.A .C., BRESLIN, A.B.X. and TERRY D.J., 1972,<br />

Asthma due to inhaled chemical agents—tolylene diisocyanate, Clin. Allergy,<br />

2. 225–36.<br />

PISATI, G., BARUFFINI, A. and ZEDDA, S., 1993, Toluene diisocyanate induced<br />

asthma: outcome according to persistence or cessation <strong>of</strong> exposure, Brit. J.<br />

Ind. Med., 50, 60–4.<br />

PROMISLOFF, R.A., PHAN, A., LENCHNER, G.S. and CICHELLI, A.V., 1990,<br />

Reactive airway dysfunction syndrome in three police <strong>of</strong>ficers following a<br />

roadside chemical spill, Chest, 98, 928–9.<br />


FRANCO, M., ZEISS, C.R. and GREENBERG, S.D., 1981, Trimellitic<br />

anhydride toxicity: a cause <strong>of</strong> acute multisystem failure, Arch. Intern. Med.,<br />

141, 1071– 4.<br />

SEGUIN, P., ALLARD, A., CARTIER, A. and MALO, J.-L., 1987, Prevalence <strong>of</strong><br />

occupational asthma in spray painters exposed to several types <strong>of</strong> isocyanates,<br />

including polymethylene polyphenylisocyanate, J. Occup. Med., 29, 340–4.<br />

VENABLES, K.M., 1989, Low molecular weight chemicals, hypersensitivity, and<br />

direct toxicity: the acid anhydrides, Brit. J. Ind. Med., 46, 222–32.<br />


antibodies against polyisocyanates in car painters, Clin. Allergy, 18, 85–93.


WERNFORS, M., NIELSEN, J. and SKERFVING, S., 1986, Phthalic<br />

anhydrideinduced occupational asthma, Int. Arch. Allergy Appl. Immunol. 79,<br />

77–82.<br />

WHITE, W.G., SUGDEN, E., MORRIS, M.J. and ZAPATA, E, 1980, Isocyanateinduced<br />

asthma in a car factory, Lancet, 1, 756–60.<br />


ROSENBERG, M. and LEVITZ, D., 1977, Trimellitic anhydride-induced<br />

airway syndromes: clinical and immunological studies, J. Allergy Clin.<br />

Immunol., 60, 96–103.


Biomarkers and risk assessment <strong>of</strong><br />

industrial chemicals

12<br />

Biomarkers and Risk Assessment<br />


Karolinska Institute, Huddinge<br />

Introduction<br />

Many chemical carcinogens cause covalent DNA-binding products,<br />

adducts, which may induce mutations or other types <strong>of</strong> DNA damage in<br />

important growth-controlling genes or loci resulting in aberrant cellular<br />

growth and cancer (Harris, 1991; IARC 1992; Hemminki, 1993). Human<br />

exposure to compounds such as polycyclic aromatic hydrocarbons (PAH)<br />

can be determined, for example, by ambient air, biological or DNA adduct<br />

monitoring. The usefulness <strong>of</strong> a method for the determination <strong>of</strong> DNA<br />

adducts in human biomonitoring requires high sensitivity because the levels<br />

<strong>of</strong> adducts are low. Here the primary focus is on the assessment <strong>of</strong><br />

exposure using the above indicators in industries where high exposure to<br />

PAHs occur, such as iron founding, coke production, aluminium<br />

production, garage work and engine overhauling with exposure to used<br />

lubricating oils.<br />

Biomonitoring <strong>of</strong> PAH exposure<br />

Literature on the application <strong>of</strong> DNA adduct studies in humans is extensive<br />

(Beach and Gupta, 1992; IARC, 1993, 1994; Hemminki et al., 1993a;<br />

Hemminki, 1994). A large majority <strong>of</strong> the 32 P-postlabelling studies on<br />

human samples focus on tobacco smoking, occupational exposures and<br />

cancer chemotherapy patients. Most occupational exposures studied relate<br />

to complex mixtures, including polycyclic aromatic hydrocarbons (PAHs).<br />

In exposure to complex mixtures multiple radioactive spots (called diagonal<br />

radioactive zones, DRZ) are detected. The adduct spots cannot be<br />

definitively identified nor quantitated. As it has turned out that for many<br />

adducts labelling is not completed, even among structural analogues such<br />

as PAHs, the adduct levels measured are likely to be underestimates<br />

(Segerbäck and Vodicka, 1993).

Table 12.1 Exposure and aromatic adducts in occupational populations, presented in simplified tabulated form<br />

K.HEMMINKI 159<br />

Notes:<br />

a Total white blood cells.<br />

b Lymphocytes.<br />

c Not given.<br />

d No data.<br />

The types <strong>of</strong> occupational groups studied by postlabelling include<br />

foundry, coke oven and aluminium workers, ro<strong>of</strong>ers, garage and terminal<br />

workers, car mechanics and chimney sweeps. All these groups have had an


increased risk <strong>of</strong> lung cancer. As, however, epidemiological studies relate to<br />

exposure a few decades earlier the risks <strong>of</strong> present exposures can only be<br />

predicted. The levels <strong>of</strong> aromatic adducts are elevated in white blood cells<br />

or lymphocytes in many <strong>of</strong> these groups. The reported total aromatic<br />

adduct levels usually range between 1 and 10 adducts per 10 8 nucleotides.<br />

Long-lived lymphocytes tend to have higher levels <strong>of</strong> adducts than shortlived<br />

granulocytes (Savela and Hemminki, 1991; Grzybowska et al., 1993).<br />

As a rule <strong>of</strong> thumb, it can be assumed that in a steady-state (i.e. long term<br />

exposure) lymphocytes contribute to the level <strong>of</strong> adducts overwhelmingly.<br />

Because they represent about 25 percent <strong>of</strong> the DNA in blood, the<br />

relationship between total white blood cell (WBC) and lymphocyte DNA<br />

adducts should be about 1:4, granulocytes only contributing to the amount<br />

<strong>of</strong> DNA denominator. Yet one has to be cautious in the comparison <strong>of</strong><br />

results between various assays even within a laboratory as the results may<br />

‘drift’ with time.<br />

The levels <strong>of</strong> white blood cell/lymphocyte aromatic adducts from<br />

workers in several industries, as measured by postlabelling, and as<br />

compared to ambient air concentrations <strong>of</strong> benzo (a) pyrene (BP) and 1hydroxypyrene<br />

levels are presented in Table 12.1. A boxplot presentation<br />

<strong>of</strong> the adduct levels <strong>of</strong> bus maintenance and truck terminal workers is<br />

shown in Figure 12.1 (Hemminki et al., 1994). The differences that were<br />

statistically significant from the controls were, in addition to the groups <strong>of</strong><br />

maintenance and terminal workers, garage workers and diesel forklift<br />

drivers.<br />

There does not seem to be a direct relationship between exposure and<br />

adduct levels. Electrode, coke and aluminium workers, exposed up to<br />

several 100 ng m −3 concentrations <strong>of</strong> BP, do not differ from the control<br />

more than foundry workers, exposed to less than 1/10 <strong>of</strong> the cited levels.<br />

The apparently higher level <strong>of</strong> adducts in the aluminium and electrode<br />

former workers (and controls) as compared to the other measurements, is<br />

due a method applied earlier with higher amounts <strong>of</strong> radioactive ATP. The<br />

later assays were carried out in small volumes but high concentrations <strong>of</strong><br />

ATP (Hemminki et al., 1993b; Szyfter et al., 1994). An increased level <strong>of</strong><br />

lymphocyte adducts has also been found in garage and truck terminal<br />

workers, with estimated exposures <strong>of</strong> about 10 ng m −3 (Hemminki et al.,<br />

1994). This would imply that the detection limit <strong>of</strong> the postlabelling<br />

method in humans exposed to PAHs lies somewhere between 1 and 10 ng<br />

m −3 BP. Whether diesel exhaust is a particularly potent inducer <strong>of</strong> adducts<br />

remains to be demonstrated. The differences between the exposed and the<br />

controls are statistically significant among foundry workers, all bus<br />

maintenance personnel and garage workers as a subgroup, all truck<br />

terminal workers and the diesel forklift drivers in particular. Coke<br />

workers differed significantly from the local controls in summer when

environmental pollution was low and the adduct levels in the controls were<br />

about 1/10 <strong>of</strong> their level in the winter (Grzybowska et al., 1993).<br />

Adducts and other endpoints<br />

K.HEMMINKI 161<br />

Figure 12.1 A boxplot <strong>of</strong> the white blood cell DNA adduct levels (per 10 8<br />

nucleotides) among bus maintenance and truck terminal workers and controls<br />

(Hemminki et al., 1994).<br />

It has become customary to include many types <strong>of</strong> endpoints to<br />

biomonitoring studies. The foundry study cited in Table 12.1 belongs to<br />

the most versatile <strong>of</strong> them. Exposure is measured by ambient air and 1hydroxypyrene<br />

monitoring (Santella et al., 1993). DNA adducts are<br />

assayed for by postlabelling and immunoassay. Plasma albumin PAH<br />

adducts are measured. Hypoxanthin guanine phosphoribosyl transferase<br />

(HPRT) and glycophorin A mutations are assayed for in lymphocytes and<br />

erythrocytes, respectively (Perera et al., 1993, 1994). Single-stand breaks in<br />

DNA and three types <strong>of</strong> cytogenetic parameters, chromosomal aberrations,<br />

sister chromatid exchanges and micronuclei, are analysed, in addition to<br />

genotyping <strong>of</strong> drug metabolising enzyme genes. Sampling <strong>of</strong> workers was<br />

repeated in four consecutive years, each at the same time <strong>of</strong> the year. As the<br />

last sampling was in the end <strong>of</strong> 1993, it will take some time before the<br />

complete data set will be available for analysis.<br />

In some published work from this data set an increase in DNA adducts<br />

and mutation frequency in the HPRT and glycophorin A genes was<br />

reported (Figure 12.2). Yet unreported results appear to show an increase


Figure 12.2 Total white blood cell DNA adducts, determined by immunoassay,<br />

HPRT and glycophorin A mutations in foundry workers, exposed to various levels<br />

<strong>of</strong> BP (Perera et al., 1993).<br />

in singlestrand breaks while none <strong>of</strong> the cytogenetic parameters are<br />

elevated in the foundry workers.<br />

Adducts and metabolic genotypes<br />

The modulation <strong>of</strong> environmental carcinogenesis by host polymorphism in<br />

genes for xenobiotics metabolising enzymes is currently under extensive<br />

investigation. It was initially sparked by findings linking certain<br />

phenotypes <strong>of</strong> drug metabolism to cancer risk (Seidegård et al., 1986;<br />

Nebert, 1991). The enzymes <strong>of</strong> interest in the context <strong>of</strong> exposure to PAHs<br />

include cytochrome P450 CYP1A1 and glutathione transferase GST,<br />

involved in the activation and inactivation, respectively, <strong>of</strong> PAHs. By<br />

restriction enzyme mapping two allelic forms, ml and m2, and two other<br />

closely linked codons for isoleucine (Ile, linked to ml) and valine (Val,<br />

linked to m2) can be defined, where m2 and valine represent the rare<br />

mutant genotypes, associated with the inducibility <strong>of</strong> the enzyme activity<br />

(Hayashi et al., 1991). Polymorphism in GST1 involves the presence or the<br />

absence <strong>of</strong> the gene (Nakachi et al., 1992). The null genotype lacks the<br />

enzyme completely.<br />

Among chimney sweeps there was an association <strong>of</strong> the rare, inducible<br />

CYP1A1 genotype ml/m2 with low adduct levels in white blood cell DNA<br />

(Ichiba et al., 1994). In the same study an increased level <strong>of</strong> adducts was<br />

noted in the GST1-individuals. The level <strong>of</strong> DNA adducts appeared to be<br />

related to both the GST and CYP1A1 genotype (Figure 12.3). Analysis <strong>of</strong><br />

micronuclei in chimney sweeps resulted in no differences between<br />

individuals <strong>of</strong> either CYP1A1 ml/m2, m2/m2 or Ile/Val genotypes nor <strong>of</strong><br />

GST1 + or − genotypes (Carstensen et al., 1993). There was however a

correlation between white blood cell DNA adducts and micronuclei and it<br />

was stronger among the GST-individuals (Ichiba et al., 1994).<br />

How important is the role <strong>of</strong> metabolic phenotype or genotype as a<br />

predictor <strong>of</strong> cancer risk remains to be established. However it would seem<br />

prudent to assume some role as long as there is significant exposure to a<br />

carcinogen, metabolism <strong>of</strong> which is regulated by polymorphic genes. It<br />

would be important to note that the question can only be addressed if both<br />

<strong>of</strong> these conditions are met. In much <strong>of</strong> the published literature there are<br />

uncertainties regarding the active agents and their metabolic routes in the<br />

tissues studied. Adjustment for a metabolic phenotype or genotype, when<br />

justified, may increase the precision in the measurement.<br />

Risk assessment<br />

K.HEMMINKI 163<br />

Figure 12.3 Total white blood cell DNA adducts, measured by postlabelling,<br />

according to CYP1A1 and GST1 genotype (Ichiba et al., 1994). Controls,<br />

Sweeps.<br />

Monitoring <strong>of</strong> DNA adducts in occupational setting has mainly been<br />

applied to workers exposed to PAHs. In the case <strong>of</strong> 32 P-postlabelling<br />

increases in the level <strong>of</strong> adducts has been noted at exposures around 10 ng<br />

BP m −3 or slightly below. This is close to the detection limit that can<br />

conveniently be attained with personal monitoring or by measuring urinary<br />

1-hydroxypyrene. As the adduct measurements also reflect some aspects <strong>of</strong><br />

metabolism and DNA repair, they extend the scope <strong>of</strong> exposure


measurements to host factors that may underly individual susceptibility to<br />

cancer.<br />

It has become increasingly common to try and incorporate other<br />

endpoints to DNA adduct studies. These include metabolic parameters,<br />

discussed above, protein adducts, cytogenetic parameters and point<br />

mutations. Examples include ethylene oxide exposed workers (Tates et al.,<br />

1991) and foundry workers (Perera et al., 1993, 1994; Santella et al.,<br />

1993). In both studies several parameters were elevated. The study on<br />

chimney sweeps illustrated how the intermediary endpoint may increase<br />

precision in the measurements (cf. Figure 12.3). The initial study showed<br />

no correlation between sweeping and micronuclei even though an<br />

adjustment was made for CYP1A1 and GST genotypes (Carstensen et al.,<br />

1993). There was a moderate correlation between sweeping and white<br />

blood cell DNA adducts, and adducts and micronuclei. Both <strong>of</strong> these<br />

correlations were strengthened once GST genotype was considered (Ichiba<br />

et al., 1994).<br />

Increasing circumstantial evidence associates DNA adducts within<br />

groups to an increased risk <strong>of</strong> cancer (IARC, 1992; Hemminki, 1993).<br />

Many <strong>of</strong> the adduct studies have been carried out in occupational groups<br />

which have been at a risk <strong>of</strong> cancer based on epidemiological results. These<br />

studies may be old and relate to exposures decades ago. Even new<br />

epidemiological publications on cancer cannot accurately address<br />

exposures after about 1970. Simultaneously there have been large changes<br />

in technology and industrial hygiene, undermining the quantitative and<br />

sometimes even the qualitative findings <strong>of</strong> the old epidemiological studies.<br />

This is one justification for the biomonitoring studies.<br />

Another justification is on exposures where epidemiological studies have<br />

not been conducted or have provided inadequate results, in spite <strong>of</strong><br />

suspicions raised by short-term or animal experiments. The International<br />

Agency for Research on Cancer has pointed out this as one <strong>of</strong> the criteria<br />

to be used in the evaluation <strong>of</strong> carcinogenicity <strong>of</strong> chemicals (IARC, 1992).<br />

Styrene belongs to this group <strong>of</strong> industrial exposures, where<br />

epidemiological findings are equivocal but adduct data are available on<br />

workers (Vodicka et al., 1993).<br />

Acknowledgements<br />

The research was supported by the Swedish Medical Research Council and<br />

Work Environment Fund.

References<br />

K.HEMMINKI 165<br />

BEACH, A. and GUPTA, R., 1992, Human biomonitoring and the 32 P-postlabeling<br />

assay, Carcinogenesis, 13, 1053–74.<br />


BRATT, I. and HAGMAR, L., 1993, B- and T-lymphocyte micronuclei in<br />

chimney sweeps with respect to genetic polymorphism for CYP1A1 and GST1<br />

(class Mu), Mutat. Res., 289, 187–95.<br />


Sea-sonal variation <strong>of</strong> aromatic DNA adducts in human lymphocytes and<br />

granulocytes, Carcinogenesis, 14, 2523–6.<br />

HARRIS, C., 1991, Chemical and physical carcinogenesis: advances and<br />

perspectives for the 1990s, Cancer Res., 51, 5023–44.<br />

HAYASHI, S., WATANABE, J., NAKACHI, K. and KAWAJIRI, K., 1991, Genetic<br />

linkage <strong>of</strong> lung cancer-associated Msp I polymorphisms with amino acid<br />

replacement in the heme binding region <strong>of</strong> the human cytochrome P4501A1<br />

gene, J. Biochem., 110, 407–11.<br />

HEMMINKI, K., 1993, DNA adducts, mutations and cancer, Carcinogenesis, 14,<br />

2007– 12.<br />

HEMMINKI, K., 1995, DNA adducts in biomonitoring, J. Occup. Environ. Med.,<br />

37, 44–51.<br />

HEMMINKI, K., AUTRUP, H. and HAUGEN, A., 1993a, Environmental<br />

carcinogens: Assessment <strong>of</strong> Exposure and Effect, pp. 89–102, Heidelberg:<br />

Springer Verlag.<br />


VODICKA, P., 1993b, Testing <strong>of</strong> quantitative parameters in the 32 Ppostlabelling<br />

method, in Phillips, D.H., Castegnaro, M. and Bartsch, H. (Eds),<br />

Postlabelling Methods for Detection <strong>of</strong> DNA Adducts, IARC Sci. Publ., No.<br />

124, pp, 51–63, Lyon: IARC.<br />


SEGERBÄCK, D., 1994, DNA adducts among personnel servicing and loading<br />

diesel vehicles, Carcinogenesis, 15, 767–9.<br />

IARC, 1992, Mechanisms <strong>of</strong> Carcinogenesis in Risk Identification, IARC Sci. Publ.<br />

No. 116, Lyon: IARC.<br />

IARC, 1993, Postlabelling Methods for Detection <strong>of</strong> DNA Adducts, IARC Sci.<br />

Publ. No. 124, Lyon: IARC.<br />

IARC, 1994, DNA Adducts: Identification and Biological Significance, IARC Sci.<br />

Publ. No. 125, Lyon: IARC.<br />


K. and HEMMINKI, K., 1994, Aromatic DNA adducts, micronuclei and<br />

genetic polymorphism for CYP1A1 and GST1 in chimney sweeps,<br />

Carcinogenesis, 15, 1347–52.<br />


HAYASHI, S.-I., WATANABE, J. and KAWAJIRI, K., 1992, High<br />

susceptibility to lung cancer analyzed in terms <strong>of</strong> combined genotypes <strong>of</strong><br />

P450IA1 and mu-class glutathione S-transferase genes, Jpn J. Cancer Res., 83,<br />



NEBERT, D.W., 1991, Role <strong>of</strong> genetics and drug metabolism in human cancer<br />

risk, Mutat. Res., 247, 267–81.<br />


1994, Biological monitoring <strong>of</strong> exposure to PAH in an electrode paste plant, J.<br />

Occup. Med., 36, 303–10.<br />


SKOGLAND, M., 1995, Studies <strong>of</strong> biomarkers in aluminium workers<br />

occupationally exposed to polycyclic aromatic hydrocarbons, Cancer<br />

Detection Prev., 19, 258.<br />



SAVELA, K. and HEMMINKI, K., 1993, HPRT and glycophorin A mutations<br />

in foundry worker: relationship to PAH exposure and PAH-DNA adducts,<br />

Carcinogenesis, 14, 969–73.<br />



K., 1994, Carcinogen-DNA adducts and gene mutations in foundry workers<br />

with changing exposure to PAH, Carcinogenesis, 15, 2905–10.<br />



and PERERA, P.P., 1993, Polycyclic aromatic hydrocarbon-DNA adducts in<br />

white blood cells and urinary 1-hydroxypyrene in foundry workers, Cancer<br />

Epidemiol. Biomarkers Prevent., 2, 59–62.<br />

SAVELA, K. and HEMMINKI, K., 1991, DNA adducts in lymphocytes and<br />

granulocytes <strong>of</strong> smokers and non-smokers detected by the 32 P-postlabelling<br />

assay, Carcinogenesis, 12, 503–8.<br />

SEGERBÄCK, D. and VODICKA, P., 1993, Recoveries <strong>of</strong> DNA adducts <strong>of</strong><br />

polycyclic aromatic hydrocarbons in the 32 P-postlabelling assay,<br />

Carcinogenesis, 14, 2463–9.<br />

SEIDEGǺRD, J., PERO, R.W., MILLER, D.G. and BEATTIE, E.J., 1986, A glutathione<br />

transferase in human leukocytes as a marker for the susceptibility to<br />

lung cancer, Carcinogenesis, 7, 751–3.<br />


HEMMINKI, K., 1994, 32 P-postlabelling analysis <strong>of</strong> DNA adducts in humans:<br />

adducts distribution and method improvement, Mut. Res., 313, 269–76.<br />




EHRENBERG, L., 1991, Biological and chemical monitoring <strong>of</strong> occupational<br />

exposure to ethylene oxide, Mut. Res., 250, 483–97.<br />

VODICKA, P., VODICKOVA, L. and HEMMINKI, K., 1993, 32 P-postlabelling <strong>of</strong><br />

DNA adduct <strong>of</strong> styrene-exposed lamination workers, Carcinogenesis, 14,<br />


13<br />

Extrapolation <strong>of</strong> Toxicity Data and Assessment<br />

<strong>of</strong> Risk<br />


Hüls AG, Marl<br />

Introduction<br />

Risk assessment provides a link between scientific research and risk<br />

management, or in other words, it is ‘a method for reaching public policy<br />

decisions’ (Silbergeld, 1993). Risk assessment includes the key elements<br />

hazard identification, dose-response assessment, and exposure assessment.<br />

These elements are integrated in a risk characterization step to predict<br />

adverse effects that may occur in a given population in a particular<br />

exposure situation, <strong>of</strong>ten based on the quantification <strong>of</strong> the likelihood <strong>of</strong> this<br />

occurrence. Risk management determines whether the particular exposure<br />

situation presents an acceptable or unacceptable risk and whether it is<br />

necessary to reduce the risk by reducing the exposure. Whereas risk<br />

management has to account for public health, socio-economical factors,<br />

technical feasibility, social perceptions, governmental policy and political<br />

consequences, risk assessment should be based on scientific principles.<br />

Since for the majority <strong>of</strong> industrial chemicals no or only limited human<br />

data exist, the question <strong>of</strong> how to extrapolate the data obtained from<br />

laboratory studies in experimental animals in order to predict the effects in<br />

humans has become one important aspect in risk assessment. The final<br />

goal is either to determine a level <strong>of</strong> exposure at which there is no reasonable<br />

doubt that an adverse effect will not occur in man or to define the risk<br />

associated with this exposure level. The use <strong>of</strong> mechanistic information to<br />

provide linkages between exposure, dose to tissue, and biological responses<br />

may assist in some <strong>of</strong> the steps necessary in the process <strong>of</strong> species<br />

extrapolation. Especially the use <strong>of</strong> physiologically based pharmacokinetic<br />

models (PBPK) for some aspects <strong>of</strong> risk assessment has been promoted to<br />

reduce the uncertainty associated with the current default methods. PBPK<br />

modeling is explained in general terms and a recently developed PBPK model<br />

for 2-butoxyethanol is provided as an example to illustrate the use <strong>of</strong><br />

kinetic and mechanistic data in risk assessment.


Legislative background in the European Union<br />

Although the process <strong>of</strong> risk assessment is not new for the individual<br />

member states <strong>of</strong> the European Union (EU) and representatives <strong>of</strong> industry<br />

and government have been assessing risk for human health and the<br />

environment for decades, EU legislation for existing 1 and new 2 substances<br />

did not formally require a systematic risk assessment up to 1992. The<br />

situation has changed with the Seventh Amendment <strong>of</strong> the Directive on the<br />

Classification, Packaging and Labelling <strong>of</strong> Dangerous Substances (EEC,<br />

1992) and the existing substances regulation (EEC, 1993a), which address<br />

risk assessment <strong>of</strong> new and existing substances, respectively.<br />

For new chemicals, the general principles <strong>of</strong> risk assessment are defined<br />

in Commission Directive 93/67/EEC (EEC, 1993b). In addition, the<br />

Directorate General XI <strong>of</strong> the European Commission has issued a series <strong>of</strong><br />

draft guidance documents for use by the competent authorities appointed<br />

by the member states. These documents provide the technical details for the<br />

risk assessment <strong>of</strong> new substances mainly by defining the testing strategies<br />

for individual toxic endpoints. For existing chemicals, a guidance<br />

document has been drafted. However, these technical guidance documents<br />

provide only little information on how to extrapolate laboratory data to<br />

humans. It has to be assumed that the extrapolation principles in the EU<br />

member states will be based on historical approaches used by authorities in<br />

other countries.<br />

Approaches to risk assessment<br />

The final goal <strong>of</strong> species extrapolation is to define a dose or dose rate<br />

which produces no adverse effects in humans. The estimation <strong>of</strong> a human<br />

no-effectlevel may include:<br />

– determination <strong>of</strong> the appropriate animal species for extrapolation to<br />

man,<br />

– determination <strong>of</strong> the most critical effect(s) and the target organ(s),<br />

– determination <strong>of</strong> the no-observed-adverse-effect level(s) (NOAEL) <strong>of</strong><br />

this effect(s), <strong>of</strong>ten followed by<br />

– extrapolation <strong>of</strong> the NOAEL(s) from subacute or subchronic to chronic<br />

exposure (time extrapolation),<br />

– extrapolation <strong>of</strong> effects observed in a high-dose region <strong>of</strong> a dose<br />

response curve to a low-dose region,<br />

1 Listed in the European Inventory <strong>of</strong> Existing Commercial Substances<br />

(EINECS).<br />

2 Not listed in EINECS.

N.FEDTKE 169<br />

– extrapolation <strong>of</strong> effects from one route <strong>of</strong> exposure to another route,<br />

and<br />

– extrapolation <strong>of</strong> effects observed in a rather homogeneous animal<br />

population to a heterogeneous human population (interspecies<br />

extrapolation), which also has to take into account the existence <strong>of</strong><br />

subgroups regarded as more sensitive as the rest <strong>of</strong> the population<br />

(intraspecies extrapolation).<br />

Essential for all procedures used in health risk assessment is the<br />

determination <strong>of</strong> the so-called critical effect. The critical effect may be<br />

defined as the adverse effect judged to be most appropriate as the basis for<br />

the risk assessment. Hence, the first step is the review <strong>of</strong> all available data<br />

on a chemical and the assessment <strong>of</strong> the adequacy <strong>of</strong> the database for the<br />

determination <strong>of</strong> the critical effect. On the basis <strong>of</strong> the critical effect,<br />

toxicants may be divided into two classes characterized by:<br />

– a threshold <strong>of</strong> response, i.e. the adverse effect on health is not expressed<br />

until the chemical, or the ultimately toxic metabolite, reaches a<br />

threshold dose or dose rate in the target tissue, or<br />

– no threshold <strong>of</strong> response, i.e. there is no threshold exposure level below<br />

which effects will not be expressed. This implies that there is some risk at<br />

any level <strong>of</strong> exposure. Examples are genotoxic carcinogens or germ cell<br />

mutagens.<br />

Based on these classes two general approaches to health risk assessment<br />

have been used.<br />

The first approach involves the use <strong>of</strong> ‘safety factors’ applied to the<br />

NOAEL or the lowest-observed-adverse-effect level (LOAEL) <strong>of</strong> a<br />

threshold effect determined in experimental animals (safety factors are<br />

recently referred to as ‘uncertainty’ or ‘assessment’ factors). The magnitude<br />

<strong>of</strong> the uncertainty factors varies between the regulatory bodies that are<br />

concerned with risk assessment, but usually they take into account the<br />

interspecies extrapolation (default factor 10) and intraspecies extrapolation<br />

(default factor 10). The magnitude <strong>of</strong> the default factors appears to be<br />

based more on the conventional use <strong>of</strong> the decimal system than on<br />

scientific reasons and have been proposed first by Lehman and Fitzhugh<br />

(1954) for the derivation <strong>of</strong> acceptable daily intakes (ADIs) for food<br />

additives. Additional uncertainty factors may be used for extrapolation to<br />

chronic exposure from subacute or subchronic exposure, adequacy <strong>of</strong> the<br />

database, extrapolation <strong>of</strong> a LOAEL to a NOAEL and severity <strong>of</strong> effects.<br />

The resulting overall uncertainty factor <strong>of</strong>ten reaches values <strong>of</strong> 1000 or<br />

higher, which is an indication <strong>of</strong> the imprecision <strong>of</strong> the derived tolerable<br />

intake. Refined extrapolation procedures using subdivisions <strong>of</strong> the default


factors or different default factors have recently been published (Lewis et<br />

al., 1990; Renwick, 1991, 1993).<br />

The second approach (for non-threshold effects) also relies mainly on<br />

default assumptions for dose-response extrapolation and cross-species<br />

extrapolation. Especially cancer risk assessment has been the subject <strong>of</strong><br />

much debate and there are a number <strong>of</strong> extrapolation methods reviewed<br />

recently by Park and Hawkins (1993) and Hallenbeck (1993). The default<br />

methodology in the . has been summarized by Frederick (1993). In<br />

principle, the risk assessment is based on a chronic rodent bioassay<br />

conducted at or near the maximum tolerated dose (MTD). The lifetime<br />

constant dose rates and the tumour incidence data for the individual dose<br />

groups are used to determine the dose response by fitting the data with a<br />

computer program. The linearized multistage cancer model (LMS) is <strong>of</strong>ten<br />

used to perform this step. The LMS model extrapolates the rodent tumor<br />

data observed at the MTD to a dose with a predefined risk and the 95 per<br />

cent upper bound on the dose-response curve is calculated. The interspecies<br />

extrapolation to humans is performed by a correction factor based on body<br />

weight or body surface. Subsequently, the dose is determined that<br />

corresponds to a maximum allowable calculated upper bound on risk. The<br />

resulting number does not describe the actual human risk under low-level<br />

environmental exposure, but provides an upper bound to human risk that<br />

is assumed not to be exceeded. The actual risk may be in the range between<br />

0 and the upper bound. In the process described, the dose is defined as<br />

administered dose or inhaled concentration. As a result, the lowdose<br />

extrapolation does not take into account non-linearities in tissue dosimetry<br />

and response. In addition, the interspecies extrapolation is performed using<br />

a default approach that does not account for mechanistic species<br />

differences.<br />

Use <strong>of</strong> PBPK models in risk assessment<br />

General description<br />

Physiologically based pharmacokinetic (PBPK) models have been used<br />

increasingly over the past decade to improve several aspects <strong>of</strong> the<br />

assessment <strong>of</strong> risk associated with human exposure to chemicals. Examples<br />

are PBPK models for styrene (Ramsey and Andersen, 1984; Csanády et al.,<br />

1994), dichloromethane (Andersen et al., 1987), 1,4-dioxane (Reitz et al.,<br />

1990a), chlor<strong>of</strong>orm (Reitz et al., 1990b), ethyl acrylate (Frederick et al.,<br />

1992), methanol (Horton et al., 1992) and 1,3-butadiene (Johanson and<br />

Filser, 1993). Recent reviews <strong>of</strong> the use <strong>of</strong> PBPK models in risk assessment<br />

have been published by several authors (Frederick, 1993; Travis, 1993;<br />

Wilson and Cox, 1993; Andersen and Krishnan, 1994).

N.FEDTKE 171<br />

PBPK models are based on the blood and tissue solubility <strong>of</strong> chemicals,<br />

their metabolism in various tissues and the physiology <strong>of</strong> the organism,<br />

thus incorporating the specific physiological description <strong>of</strong> animal species as<br />

well as specific physico-chemical descriptions <strong>of</strong> agents. Uptake,<br />

distribution, metabolism and excretion are described in physiologically<br />

realistic compartments (tissue groups) using computer simulation. The<br />

compartments are linked in parallel, represent the actual mammalian<br />

architecture, and include tissues such as lung and arterial blood, fatty<br />

tissue, poorly perfused tissues (muscles, skin), richly perfused tissues<br />

(brain, kidneys, heart, endocrine gland, gastro-intestinal tract), liver as the<br />

main metabolizing tissue, and mixed venous blood. The compartments are<br />

connected by arterial and venous blood flow and are characterized by a set<br />

<strong>of</strong> mass balance differential equations. The rate constants that describe the<br />

flow <strong>of</strong> material between the tissue groups and the rate <strong>of</strong> change in the<br />

chemical concentration <strong>of</strong> each compartment are proportional to blood<br />

flow, tissue solubility and compartment volumes. The basic mathematical<br />

description <strong>of</strong> a PBPK model for a volatile compound has been provided by<br />

Ramsey and Andersen (1984) and additional details may be found in<br />

appendices <strong>of</strong> manuscripts dealing with the development <strong>of</strong> PBPK models.<br />

Estimation <strong>of</strong> the constants used in PBPK models may be based on the<br />

literature in the case <strong>of</strong> physiological parameters such as ventilation rates,<br />

cardiac output, blood flow to tissues and tissue volumes. The EPA has<br />

compiled reference values for these parameters and their scaling (Arms and<br />

Travis, 1988). Chemical specific parameters such as blood and tissue<br />

solubilities may be determined from in vitro preparation (Sato and<br />

Nakajima, 1979; Gargas et al., 1989). The biochemical constants for<br />

metabolism may be derived from in vitro studies (Reitz et al. 1988;<br />

Carfagna and Kedderis, 1992; Johanson and Filser, 1993), in vivo<br />

toxicokinetic studies (Potter and Tran, 1993; Frederick et al., 1992) or in<br />

the case <strong>of</strong> volatile substances from gas uptake studies (Gargas et al., 1986,<br />

1990; Filser, 1992).<br />

Since the tissue groups have a defined biological meaning, scaling <strong>of</strong> the<br />

associated parameters between species is possible since many <strong>of</strong> the<br />

parameters used are correlated to body weight. Cardiac output, alveolar<br />

ventilation rate and V max are scaled by the 3/4 power <strong>of</strong> body weight<br />

whereas K m is assumed to be constant across species. However, the<br />

substitution <strong>of</strong> the physiological parameters with the appropriate values<br />

characteristic for the species <strong>of</strong> interest is preferred.<br />

The development <strong>of</strong> PBPK models is an iterative process involving<br />

comparison <strong>of</strong> the model simulations with experimental data and<br />

refinement <strong>of</strong> the estimates when the model fails to accurately predict the<br />

kinetic behaviour. Different exposure scenarios can be used to predict the<br />

concentrations <strong>of</strong> the parent chemical or its metabolites in the blood or the<br />

tissues, which are the target <strong>of</strong> toxic effects. The level <strong>of</strong> glutathione


depletion in hepatic and extrahepatic tissues (D’Souza et al., 1988;<br />

Frederick et al., 1992; Krishnan et al., 1992), kinetic interactions <strong>of</strong> parent<br />

compounds in mixed exposures (Tardif et al., 1993) or the amount <strong>of</strong><br />

adducts formed by macromolecular binding (Krishnan et al., 1992) are<br />

predictions that may also be generated by PBPK modeling. As a result <strong>of</strong><br />

the simulations, quantitative information on the internal dose <strong>of</strong> a<br />

chemical or its metabolites in the target tissue is obtained and can replace<br />

the administered dose conventionally used in risk assessment. After<br />

validation <strong>of</strong> the PBPK models in experimental animals, human PBPK<br />

models can be developed either by allometric scaling <strong>of</strong> the physiological<br />

and biochemical parameters or preferably using the actual human<br />

parameters. Following the prediction <strong>of</strong> the target tissue dosimetry in<br />

humans, the appropriate dose surrogates are related to the effect <strong>of</strong> interest<br />

and quantitative species differences are determined. This information<br />

provides the possibility to base the species extrapolation on scientific data<br />

instead <strong>of</strong> on arbitrarily assigned default factors and as a consequence the<br />

uncertainty <strong>of</strong> the extrapolation procedures applied in conventional risk<br />

assessment may be reduced.<br />

Description and use <strong>of</strong> the PBPK model for 2butoxyethanol<br />

2-Butoxyethanol (BE) is a widely produced glycol ether used as a key<br />

ingredient in water- or solvent-based coatings, industrial and consumer<br />

cleaning products, and as solvent in a variety <strong>of</strong> products. Haemolysis was<br />

identified as most sensitive indicator <strong>of</strong> BE-induced toxicity in several<br />

species <strong>of</strong> laboratory animals and has received the most attention as a<br />

critical effect for human risk assessment (ECETOC, 1985, 1994). The<br />

experimentally determined subchronic NOAEL for the rat is 25 ppm. The<br />

major metabolite <strong>of</strong> BE is 2-butoxyacetic acid (BAA) which has been<br />

identified as the metabolite responsible for the haemolysis <strong>of</strong> red blood<br />

cells in in vitro and in vivo studies (Bartnik et al., 1987; Ghanayem et al.,<br />

1987; Ghanayem, 1989). Changes in the deformability <strong>of</strong> rat erythrocytes<br />

appear to precede haemolysis upon treatment with BAA. Treatment <strong>of</strong><br />

human erythrocytes with BAA did not induce changes in deformability<br />

(Udden and Patton, 1994; Udden, 1994). The observed species differences<br />

may be due to differences in the lipid composition <strong>of</strong> erythrocyte<br />

membranes, differences in membrane proteins associated with anion<br />

transport processes, or differences in the erythrocyte cytoskeleton (Udden,<br />

1994; Udden and Patton, 1994). Humans are most likely to be exposed to<br />

BE by the dermal or inhalation routes due to the widespread use <strong>of</strong> BE in<br />

cleaning products. Assessment <strong>of</strong> the risk resulting from BE use has to<br />

account for these routes <strong>of</strong> exposure and the formation <strong>of</strong> BAA as the<br />

active metabolite. In order to assist in the risk assessment, PBPK models

N.FEDTKE 173<br />

were developed that describe the uptake, metabolism and disposition <strong>of</strong> BE<br />

and BAA (Johanson, 1986; Corley et al., 1993, 1994; Shyr et al., 1993).<br />

The model <strong>of</strong> Corley et al. (1993, 1994) is a refinement <strong>of</strong> Johanson’s<br />

model (1986) and consists <strong>of</strong> two submodels. The first submodel describes<br />

the uptake and disposition <strong>of</strong> BE and consists <strong>of</strong> the tissue compartments<br />

rapidly perfused organs, slowly perfused organs, fat, skin, muscle,<br />

gastrointestinal tract, and liver as the metabolizing tissue. The BE<br />

submodel allows uptake via the inhalation and dermal routes and in<br />

addition provides the possibility <strong>of</strong> uptake via IV infusion and the<br />

gastrointestinal tract in order to validate the model with laboratory data.<br />

The second submodel tracks the disposition <strong>of</strong> BAA in the same tissue<br />

compartments, but the kidney was removed from the rapidly perfused<br />

organs as separate tissue to allow for the excretion <strong>of</strong> BAA metabolites.<br />

The two submodels are linked together by the metabolism <strong>of</strong> BE to BAA<br />

via a saturable enzymatic pathway catalyzed by alcohol and aldehyde<br />

dehydrogenases in the liver. Competing pathways (BE conjugation and<br />

BE O-dealkylation) are lumped together and described by an additional<br />

enzymatic pathway with Michaelis-Menten kinetics. The model assumes<br />

that BAA is bound to proteins in blood and is eliminated by a saturable<br />

process in the kidneys. The rate <strong>of</strong> BAA elimination by the kidneys is<br />

described as the sum <strong>of</strong> glomerular filtration rate <strong>of</strong> BAA and the acid<br />

transport <strong>of</strong> BAA assuming that no reabsorption occurs. The biochemical<br />

constants determined experimentally in the rat were scaled to humans by<br />

(body weight) 0.7 . In the validation process, the model successfully described<br />

a wide variety <strong>of</strong> rat and human data from different laboratories using<br />

several routes <strong>of</strong> administration.<br />

BAA was predicted to be formed more rapidly in rats compared with<br />

humans, but to be eliminated slower in humans than in rats. In summary,<br />

higher maximum concentrations <strong>of</strong> BAA in blood (C max) and also higher<br />

areas under the BAA concentration-time curves (AUC) were predicted for<br />

rats than for humans, especially as the vapour concentration was<br />

increased. For the purpose <strong>of</strong> dose-response and interspecies extrapolation,<br />

BAA-C max and BAA-AUC were used as estimates <strong>of</strong> the internal dose<br />

surrogate; C max can be related directly to the in vitro haemolysis studies<br />

with BAA and is responsive to the dose-rate. The in vitro studies performed<br />

(Bartnik et al., 1987; Ghanayem et al., 1987; Ghanayem, 1989; Udden,<br />

1994; Udden and Patton, 1994) suggest that approximately 0.2 mM BAA<br />

is required to produce slight haemolysis <strong>of</strong> rat red blood cells. At about 2<br />

mM BAA nearly complete haemolysis was observed. The model predicts<br />

for nose-only exposure that these concentrations are reached in the rat at<br />

BE exposure concentrations <strong>of</strong> about 100 ppm and 800 ppm for 6 h,<br />

respectively, which is consistent with observations in vivo (Tyler, 1984;<br />

Sabourin et al., 1992).


For human red blood cells, the minimum BAA-concentration necessary<br />

to induce slight haemolysis is about 40 times higher compared with rats,<br />

i.e. 8 mM. The model predicts for human nose-only exposure that the C max<br />

<strong>of</strong> BAA in blood is slightly lower than the value observed in rats at a BE<br />

exposure concentration <strong>of</strong> about 100 ppm for 6 h and is only about 50 per<br />

cent <strong>of</strong> the BAA rat blood concentration at 800 ppm. In any case, the<br />

minimum toxic concentration <strong>of</strong> approximately 8 mM BAA in human<br />

blood is not achieved.<br />

The AUC has a time component which is important since haemolysis is<br />

not an instantaneous response (Udden, 1994; Udden and Patton, 1994).<br />

With respect to the AUCs for BAA, the model predicts that the values for<br />

rat and human blood are similar up to a BE exposure concentration <strong>of</strong><br />

about 500 ppm. Higher BE concentrations cause higher AUCs for BAA in<br />

rat blood than in human blood. Thus, the model predicted a BAA-AUC in<br />

man at 22 ppm BE vapour exposure that was similar to the BAA-AUC in<br />

rats achieved at 25 ppm BE vapour exposure, the established subchronic<br />

NOAEL.<br />

The simulation <strong>of</strong> the dermal BE uptake assumed that 10 per cent <strong>of</strong> the<br />

body surface <strong>of</strong> rats and humans were exposed for 6 h to BE solutions in<br />

water (5–100 per cent) and that no losses <strong>of</strong> BE occurred from the dosing<br />

solution. The simulation predicted C max blood concentrations for BAA in<br />

rats that were highest (about 3 mM) for a 40 per cent BE solution. For<br />

humans, BAA-C max was predicted to reach about 1.3 mM for the same BEconcentration.<br />

Predicted BAA-AUCs were about tw<strong>of</strong>old higher in rats<br />

compared with humans. Under these worst-case assumptions, no BE<br />

concentration is expected to achieve BAA concentrations in human blood<br />

that would cause haemolysis.<br />

ECETOC (1994) used the described PBPK model for BE and BAA<br />

disposition in combination with mechanistic data obtained by in vitro<br />

experiments to recommend an occupational exposure limit for BE:<br />

– BAA-AUC was used as the internal dose surrogate and 22 ppm BE<br />

vapour was predicted to cause a BAA-AUC in human blood similar to<br />

the BAA-AUC in rats exposed to a BE-concentration <strong>of</strong> 25 ppm<br />

(identified as the subchronic rat NOAEL). At 22 ppm BE vapour, the<br />

BAA-C max (33 µM) is predicted to be several hundredfold below the<br />

BAA concentration that causes pre-haemolytic effects in human red<br />

blood cells (8 mM).<br />

– ECETOC did not use an uncertainty factor for intraspecies<br />

extrapolation, since the in vitro studies indicated no increased sensitivity<br />

<strong>of</strong> red blood cells from individuals regarded as susceptible to haemolytic<br />

effects such as older persons, persons with hereditary spherocytosis or<br />

sickle cell disease (Udden, 1994; Udden and Patton, 1994).

– An uncertainty factor for time extrapolation (subchronic to chronic<br />

exposure) was also not applied, since the red blood cell haemolysis was<br />

regarded as a transient phenomenon observed predominantly on the<br />

first few days <strong>of</strong> exposure thus indicating that longer exposure would<br />

not have resulted in a lower rat NOAEL.<br />

– Although there is some uncertainty about the actual magnitude <strong>of</strong> the<br />

contribution <strong>of</strong> dermal uptake to the total uptake during BE vapour<br />

exposure, ECETOC concluded that even under worst-case conditions<br />

the BAA concentrations achieved are not sufficient to cause haemolysis<br />

in man and there is no need for the adjustment <strong>of</strong> the predicted human<br />

NOAEL for route.<br />

In conclusion, an occupational exposure limit <strong>of</strong> 20 ppm (8 h TWA) was<br />

recommended, also taking into account all other effects that may be<br />

associated with BE-exposure. This value is similar to the rat NOAEL <strong>of</strong> 25<br />

ppm for the most sensitive parameter, i.e. haemolysis, and was derived<br />

using scientific data instead <strong>of</strong> applying default factors to the rat NOAEL,<br />

a procedure which would have overpredicted the human risk associated<br />

with BE-exposure.<br />

Conclusion<br />

N.FEDTKE 175<br />

The use <strong>of</strong> PBPK models and mechanistic data in risk assessment tends to<br />

reduce the uncertainties in comparison with default methodologies by<br />

replacing the administered dose with the delivered dose and also tends to<br />

reveal uncertainties concealed in default methodologies (Wilson and Cox,<br />

1993). However, there are also limitations in the development <strong>of</strong> PBPK<br />

models. One limitation is that the mechanism <strong>of</strong> the toxic effect has to be<br />

known, otherwise the replacement <strong>of</strong> the external dose by internal dose<br />

surrogates is not possible. In addition, extensive validation <strong>of</strong> the model is<br />

necessary in order to replace default approaches in risk assessment. For the<br />

time being, the development <strong>of</strong> PBPK models appears to be restricted to<br />

high production chemicals where the existing data base allows<br />

identification <strong>of</strong> an accepted mechanism <strong>of</strong> toxic action and validation <strong>of</strong><br />

the model. Concern has been expressed that the use <strong>of</strong> point estimates in<br />

PBPK modelling instead <strong>of</strong> ranges <strong>of</strong> biologically plausible values leads to<br />

an increase in the uncertainty (Portier and Kaplan, 1989). However, a<br />

recent study from the Delivered Dose Work Group <strong>of</strong> the American<br />

<strong>Industrial</strong> Health Council came to the conclusion that incorporation <strong>of</strong><br />

‘pharmacokinetic information in a risk assessment,…, leads to both a more<br />

accurate estimate <strong>of</strong> risk and a better specification <strong>of</strong> the true uncertainty’<br />

(Wilson and Cox, 1993). A detailed discussion <strong>of</strong> the sources <strong>of</strong><br />

uncertainties is also provided in this reference.


If the data base is sufficient, PBPK models provide scientific credibility to<br />

interspecies extrapolation, extrapolation across routes <strong>of</strong> administration,<br />

extrapolation from high-dose to low-dose and intraspecies extrapolation.<br />

Recent concepts link the original tissue dose concept <strong>of</strong> PBPK models to<br />

biologically based tissue response models, thus relating the delivered dose<br />

via the mechanism <strong>of</strong> action to the toxic response and developing<br />

integrated biological models (Conolly et al., 1988; Moolgavker et al.,<br />

1988; Cohen and Ellwein, 1990; Conolly and Andersen, 1991). Such<br />

approaches enable scientists to ask the right questions and to design new<br />

mechanistic studies that will lead toward the goal <strong>of</strong> a scientifically-based<br />

risk assessment.<br />

Acknowledgement<br />

The author thanks Richard A.Corley and the Glycol Ether Panel <strong>of</strong> the<br />

Chemical Manufactures Association for providing data on 2butoxyethanol.<br />

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14<br />

Molecular Approaches to Assess Cancer Risks<br />



Sittingbourne Research Centre, Sittingbourne<br />

Introduction<br />

Carcinogenesis is a complex process which is not yet fully understood.<br />

Nevertheless, it is generally accepted that carcinogenesis involves the<br />

accumulation <strong>of</strong> mutations in critical genes: proto-oncogenes and/or<br />

tumour suppressor genes. These mutations transform normal cells into<br />

‘initiated’ cells possessing the full complement <strong>of</strong> genetic changes necessary<br />

for malignancy (Figure 14.1). The critical mutations may result from<br />

exposures to radiation, to genotoxic chemicals or they may arise<br />

‘spontaneously’ as a consequence <strong>of</strong> miscoding errors during the normal<br />

replication <strong>of</strong> DNA. Concomitantly, mutations will also accumulate in<br />

other genes which, although not critical for cancer per se may,<br />

nevertheless, influence cellular character thereby contributing to the<br />

multifaceted nature <strong>of</strong> cancer.<br />

The precise nature and number <strong>of</strong> critical genetic changes required for<br />

initiation have not yet been established but will probably vary from case to<br />

case. Many researchers envisage a strict temporal sequence <strong>of</strong> genetic<br />

changes in carcinogenesis. However, it is probable that the critical<br />

mutations can occur in any sequence and at any time. Indeed, it is clear<br />

that one or more <strong>of</strong> the critical mutations can occur in parental cells.<br />

Transmission (inheritance) <strong>of</strong> these mutations either through the germ line<br />

or via somatic cell division increases the susceptibility <strong>of</strong> the progeny to<br />

carcinogens. Furthermore, it is important to note that each <strong>of</strong> the critical<br />

mutations necessary for malignancy may have a different cause. This<br />

potential for multiple causation has important implications in risk<br />

assessment (vide infra).<br />

Fully initiated cells may not automatically proliferate to form tumours.<br />

One possible explanation is that the surrounding normal cells restrain<br />

the initiated or latent cancer cell by providing essential growth regulators<br />

which are no longer produced by the initiated cell. Nevertheless, partially<br />

and fully initiated cells have a replicative and/or survival advantage over<br />

normal cells. Tissue injury caused by physical trauma, chemical agents or

Figure 14.1 Schematic representation <strong>of</strong> the carcinogenic process.<br />

A.S.WRIGHT ET AL. 181<br />

viruses may have a derestraining effect thereby triggering or facilitating the<br />

replication <strong>of</strong> partially or fully initiated cells to form benign tumours or<br />

malignant tumours. Increased functional demands may also serve to<br />

promote tumour development in affected tissues.<br />

It is clear that chemicals which promote tumour development are very<br />

important determinants <strong>of</strong> carcinogenesis. Indeed, promoting agents<br />

display a marked tendency for organotropism. Promoter action is,<br />

therefore, probably the most important determinant <strong>of</strong> the site <strong>of</strong> tumour<br />

development. Yet genotoxic chemicals which initiate the carcinogenic<br />

process are perhaps viewed with even greater concern. The reasons for this<br />

high level <strong>of</strong> concern hinge mainly on evidence that the mutagenic or<br />

initiating actions <strong>of</strong> genotoxic chemicals are additive, cumulative and<br />

essentially irreversible. Furthermore, in contrast to most other classes <strong>of</strong><br />

toxic chemicals, including promoters operating via cytotoxic mechanisms,<br />

there is no theoretical reason or experimental evidence to support the view<br />

that mutagenic actions <strong>of</strong> genotoxic chemicals are thresholded. For these<br />

reasons even very low exposures <strong>of</strong> genotoxic chemicals are viewed with<br />

concern. These concerns have focused scientific and regulatory attention on<br />

a need to develop sound approaches to manage cancer risks—particularly<br />

low level risks associated with low exposures to genotoxic chemicals<br />

encountered in the occupational or environmental settings. Indeed, apart<br />

from clinical applications, high exposures to genotoxic chemicals cannot be<br />



Management <strong>of</strong> cancer risks (key requirements)<br />

The management <strong>of</strong> toxicological risks implies a capacity to control<br />

exposures within acceptable safety limits. Effective control is, therefore,<br />

dependent not only on the qualitative detection and identification <strong>of</strong><br />

hazardous chemicals but also on a capacity to determine human exposure<br />

and to evaluate the health risks. This last requirement necessitates a<br />

knowledge <strong>of</strong> potency, i.e. quantitative human dose-response relationships.<br />

In the case <strong>of</strong> genotoxic chemicals, the relevant data reside in the very low<br />

region <strong>of</strong> the dose-response curve.<br />

The concept <strong>of</strong> acceptable risk is readily accepted when applied to the<br />

many classes <strong>of</strong> toxic chemicals operating by a thresholded mechanism<br />

indicative <strong>of</strong> the virtual absence <strong>of</strong> risk at sub-threshold doses. Absolute<br />

safety margins for genotoxic chemicals cannot be guaranteed (vide supra)<br />

leading to the adoption <strong>of</strong> conservative safety measures. Indeed, cursory<br />

analysis might suggest that quantitative risk data are not required for the<br />

effective management <strong>of</strong> cancer risks associated with genotoxic chemicals.<br />

Thus, it is generally accepted that human contact with carcinogens should<br />

be minimised. Purely qualitative identification <strong>of</strong> the hazard would permit<br />

the design <strong>of</strong> measures to limit human exposure and minimise carcinogenic<br />

impact. Indeed, a purely qualitative indication <strong>of</strong> genotoxicity can be an<br />

absolute deterrent to the development <strong>of</strong> new products. Nevertheless,<br />

certain exposures, e.g. to indigenous genotoxic chemicals and natural food<br />

components, are unavoidable. Furthermore, measures to reduce exposures<br />

to ‘avoidable’ genotoxic hazards, e.g. certain combustion products, key<br />

industrial base chemicals and intermediates, are <strong>of</strong>ten difficult and costly.<br />

Quantitative risk assessment is needed to prioritise these hazards and, most<br />

importantly, to determine safety margins. Certainly a failure to determine<br />

carcinogenic potency would lead to uncertainty about the adequacy <strong>of</strong><br />

safety margins and, probably, to unnecessary measures to further reduce<br />

exposures. Thus, despite their additive and cumulative actions, even<br />

genotoxic chemicals can pose a negligible health risk. Of course, the<br />

definition <strong>of</strong> a negligible, i.e. acceptable, risk is a socio-political judgement<br />

which nevertheless has to be realistic in the case <strong>of</strong> unavoidable hazards<br />

and achievable in the case <strong>of</strong> avoidable hazards.<br />

Detection <strong>of</strong> genotoxic carcinogens<br />

Concerns about genotoxic hazards have provided an incentive for the<br />

development <strong>of</strong> a broad range <strong>of</strong> rapid tests to detect intrinsic genotoxic<br />

activity or potential. The principal aim <strong>of</strong> these approaches is to predict<br />

carcinogenic activity or, more accurately, cancer initiating activity. The<br />

most widely used tests are the coupled microsomal-microbial mutation<br />

assays developed by Ames et al. (1973). However, such approaches are

A.S.WRIGHT ET AL. 183<br />

viewed as too remote to be <strong>of</strong> value in estimating cancer risks. The trend is<br />

towards increasingly sensitive and precise technology—particularly generic<br />

methods with potential for direct application in humans.<br />

Advances in molecular biology have permitted the development <strong>of</strong> a new<br />

generation <strong>of</strong> point mutation assays based on DNA base mismatch<br />

technology (Thilly, 1991; Lu and Hsu, 1992). This technology has a<br />

precision far exceeding that <strong>of</strong> conventional biological methods and a<br />

sensitivity permitting direct applications in humans. The full potential <strong>of</strong> this<br />

technology has not yet been realised. However, it seems probable that<br />

detection levels will ultimately obviate a need for prior phenotypic<br />

selection: paving the way to universal application. Avoidance <strong>of</strong> phenotypic<br />

selection would represent a powerful advantage over existing<br />

methodologies by providing a much more direct and reliable route to<br />

determining overall background mutation rates and increments due to<br />

specific exposures <strong>of</strong> key relevance to cancer risk assessment (vide infra).<br />

The most prospective <strong>of</strong> the current assays are those designed to detect<br />

primary DNA damage. Among these procedures, 32 P-post-radiolabelling<br />

technology developed by Randerath et al. (1981) to detect DNA adducts is<br />

by far the most sensitive. The justification for application <strong>of</strong> such a<br />

prospective approach to detect exposure to genotoxic carcinogens hinges<br />

on the causal relationship established between genotoxic activity and<br />

cancer. In general, genotoxic character is conferred by possession <strong>of</strong> a<br />

centre(s) <strong>of</strong> electrophilic reactivity. This reactivity permits the chemical to<br />

undergo chemical reactions with nucleophilic centres in the target molecule<br />

(DNA). In many instances the electrophilic centre(s) is introduced into an<br />

inactive precursor chemical by metabolic activation. Primary products, e.g.<br />

DNA adducts, formed when genotoxic chemicals react with DNA are<br />

generally promutagenic (or lethal) and their occurrence leads to an<br />

increased risk <strong>of</strong> mutation and cancer. There is no known category <strong>of</strong><br />

chemical which forms DNA adducts that can be excluded from this<br />

generalisation. Not all DNA adducts are strongly promutagenic. However,<br />

because electrophiles do not display absolute specificity in their reactions<br />

with nucleophiles, the detection <strong>of</strong> even a weakly promutagenic adduct,<br />

e.g. N 7 -alkyldeoxyguanosine, signals the formation <strong>of</strong> a more strongly<br />

promutagenic adduct, e.g. O 6 -alkyldeoxyguanosine. If follows that the<br />

detection <strong>of</strong> DNA adducts provides qualitative evidence <strong>of</strong> (human)<br />

exposure to a genotoxic carcinogen.


Identification <strong>of</strong> human carcinogens<br />

Classical epidemiological approaches<br />

Until very recently, epidemiological approaches to detect and identify<br />

environmental carcinogens were based exclusively on the analysis <strong>of</strong><br />

tumour incidence and chromosome aberrations in human populations.<br />

However, the endpoints <strong>of</strong> these biological methods lack the intrinsic<br />

resolving power needed to dis criminate between different contributory<br />

factors. Indeed, it is only in instances <strong>of</strong> specific, high and, <strong>of</strong>ten, localised<br />

exposures that these methods have been effective in identifying specific<br />

causative agents. Nevertheless, the results <strong>of</strong> epidemiological studies<br />

indicate that chemicals, which may include both natural and xenobiotic<br />

compounds in food, drink or in the local or general environment, play a<br />

major and broad role in the aetiology <strong>of</strong> human cancer. The identification<br />

<strong>of</strong> these chemical factors is a major goal in cancer prevention.<br />

In vitro genotoxicity assays<br />

In addition to applications in screening prospective chemical products, in<br />

vitro genotoxicity assays, particularly the Ames test, provided the first<br />

practicable, systematic approach to identify environmental carcinogens.<br />

However, this approach places very heavy demands on the time and effort<br />

required to fractionate environmental samples and test individual<br />

compounds. More importantly, however, like the animal cancer studies<br />

these assays complement or have largely supplanted, the approach is not<br />

specifically targeted towards identifying and prioritising human hazards.<br />

For example, these short-term in vitro test do not provide direct evidence<br />

<strong>of</strong> human exposure or effects.<br />

Molecular epidemiology<br />

32 P-Post-radiolabelling technology for the analysis <strong>of</strong> DNA adducts<br />

provides the basis <strong>of</strong> a very sensitive and generic approach to detect<br />

exposures to genotoxic carcinogens. This technology has universal<br />

application and can be applied to detect DNA adducts formed in<br />

laboratory species or humans during exposures to both known and, as yet,<br />

unidentified genotoxic chemicals at the low concentrations encountered in<br />

the environment and the workplace. Elucidation <strong>of</strong> the chemical structures<br />

<strong>of</strong> adducts in human DNA would provide a basis for identifying the<br />

causative agents and their sources or origins. This possibility <strong>of</strong> identifying<br />

the chemical initiators <strong>of</strong> human cancer is an exciting prospect.<br />

Unfortunately, however, these adducts are present at very low abundances<br />

and this is a major obstacle to identification. Thus, the methods for

detecting DNA adducts are much more sensitive than the physicochemical<br />

methods needed for structural characterisation. A number <strong>of</strong> strategies<br />

have been adopted in attempts to solve this problem.<br />

Protein adducts<br />

Genotoxic chemicals that react with DNA also react with nucleophilic<br />

centres in proteins and may also undergo ‘spontaneous’ and enzymecatalysed<br />

reactions with glutathione leading to the excretion <strong>of</strong> the<br />

corresponding mercapturic acids. Ins<strong>of</strong>ar as the formation <strong>of</strong> protein<br />

adducts and mercapturic acids reflect the formation <strong>of</strong> the corresponding<br />

DNA adducts, their detection may also furnish evidence <strong>of</strong> exposure to a<br />

genotoxic carcinogen.<br />

The potential for reaction <strong>of</strong> genotoxic chemicals with proteins (and<br />

glutathione) is much greater than with DNA. Furthermore, human<br />

proteins, e.g. haemoglobin, are available in much larger quantities and are<br />

more accessible than human tissue DNA. These advantages have been<br />

exploited, particularly in the pioneering work <strong>of</strong> Ehrenberg’s group, to<br />

develop a range <strong>of</strong> procedures for the qualitative and quantitative analysis<br />

<strong>of</strong> protein adducts (Osterman-Golkar et al., 1976; Calleman et al., 1978;<br />

Ehrenberg and Osterman-Golkar, 1980). (For a review <strong>of</strong> the available<br />

methods see Skipper and Naylor, 1991.) The most powerful and generic<br />

approach is undoubtedly that developed by Törnqvist et al. (1986a). An<br />

initial purification or enrichment step is key to any successful method for<br />

the analysis <strong>of</strong> low levels <strong>of</strong> organic residues. The amino-groups <strong>of</strong> the Nterminal<br />

valine residues <strong>of</strong> the α-and<br />

β-chains <strong>of</strong> human haemoglobin are<br />

major targets for reaction with a broad range <strong>of</strong> genotoxic chemicals.<br />

Törnqvist achieved selective enrichment <strong>of</strong> adducted N-terminal valine<br />

residues <strong>of</strong> haemoglobin by devising a modified Edman degradation which<br />

resulted in the scission <strong>of</strong> adducted residues whilst leaving the nonadducted<br />

N-terminal valines intact. This procedure provides the basis for<br />

identifying the adducting moieties and their quantitation by GC/MS.<br />

Applications <strong>of</strong> this technology have furnished evidence <strong>of</strong> background<br />

exposures to a range <strong>of</strong> alkylating species. Protein adduct technology has<br />

the potential for considerable further refinement. The possibility <strong>of</strong> using<br />

immunoaffinity technology to enrich both known and unidentified protein<br />

adducts is currently being explored.<br />

DNA adducts and immunoenrichment<br />

A.S.WRIGHT ET AL. 185<br />

The need for effective enrichment technology for DNA adducts is even<br />

more pressing than for protein adducts. Ideally, the enrichment procedure<br />

should be applied at the earliest possible stage <strong>of</strong> analysis. The procedure


should be rapid and mild in order to minimise the formation <strong>of</strong> artefacts.<br />

Currently, immunoaffinity technology holds the greatest promise.<br />

Immunoenrichment <strong>of</strong> DNA adducts necessitates antibodies possessing<br />

the appropriate specificities and affinities to permit selective binding <strong>of</strong><br />

adducts at the very low abundances encountered in hydrolysates or enzyme<br />

digests <strong>of</strong> human DNA. The immune system does not normally respond to<br />

small molecules per se. However, the system can be induced to produce<br />

effective antibodies by immunising with the small molecule (hapten)<br />

coupled to a protein. Such treatment induces a spectrum <strong>of</strong> antibodyproducing<br />

cells, each producing a specific antibody. Most <strong>of</strong> these<br />

antibodies recognise various regions (epitopes) <strong>of</strong> the carrier protein while<br />

a few may specifically recognise and bind the small molecule <strong>of</strong> interest.<br />

Suitable antibody-producing cells can be selected and cloned to provide a<br />

permanent source <strong>of</strong> homogenous antibody (monoclonal antibody, Mab).<br />

Mabs can be raised against virtually any organic chemical although some<br />

lower molecular weight compounds (

adducts are not viewed as particularly promutagenic. Nevertheless the N 7 -<br />

atom <strong>of</strong> dG residues in DNA is a major target for adduction and the<br />

detection <strong>of</strong> N 7 -dG adducts signals the production <strong>of</strong> adducts at other<br />

(more critical) sites in DNA (vide supra).<br />

In certain instances, a class <strong>of</strong> adducting moieties may possess a common<br />

structural feature that can be exploited for immunoenrichment. For<br />

example, a Mab has been raised against the major DNA adduct <strong>of</strong> benzo(a)<br />

pyrene (r-7,t-8,t-9-trihydroxy-c-10-(N 2 -deoxyguanosylphosphate)-7,8,9,<br />

10-tetrahydrobenzo(a)pyrene) in a collaborative study with Dr Baan’s<br />

group. This Mab recognises DNA adducts formed by a broad range <strong>of</strong><br />

polycyclic aromatic hydrocarbons (PCAs) including benzo(a)pyrene (BP),<br />

chrysene, benz(a)anthracene, 5-methylchrysene, picene and dibenz(a,h)<br />

anthracene. It seems probable that this Mab recognises the common<br />

trihydric alcohol structure produced when the reactive diol epoxides <strong>of</strong><br />

each <strong>of</strong> these polycyclic compounds reacts with nucleophilic centres in<br />

DNA or other macromolecules. The fact that the Mab does not bind the<br />

corresponding fluoranthene adduct is consistent with the spatial<br />

environment <strong>of</strong> the hydroxyl groups in fluoranthene-DNA adducts which<br />

is completely different from those generated from the other PCAs employed<br />

in this study.<br />

The performance <strong>of</strong> the Mab raised against the major BP-DNA adduct in<br />

the enrichment <strong>of</strong> PCA-DNA adducts is being evaluated using the<br />

immobilised Mab coupled to cyanogen bromide-activated Sepharose 4B.<br />

Results obtained to date demonstrate that the immobilised Mab selectively<br />

adsorbs the major BP-DNA adduct from DNA hydrolysates at abundances<br />

below 1 adduct per 10 9 nucleotide units. Results with the other PCA-DNA<br />

adducts are not yet available. However, the results obtained with the major<br />

BP-DNA adduct underlines the potential <strong>of</strong> immunoenrichment technology<br />

in the qualitative and quantitative analysis <strong>of</strong> adducts. Furthermore, such<br />

results provide an incentive to pursue the development <strong>of</strong> class-specific<br />

antibodies in order to permit or facilitate the identification <strong>of</strong> the chemical<br />

initiators <strong>of</strong> human cancer.<br />

Mercapturic acids<br />

A.S.WRIGHT ET AL. 187<br />

Qualitative analysis <strong>of</strong> mercapturic acids also provides a basis for<br />

identifying human exposures to genotoxic chemicals (vide supra). However,<br />

the available analytical procedures are complex and tend to lack specificity<br />

and sensitivity. During the last dozen years we have undertaken a number<br />

<strong>of</strong> studies aimed at developing compound- and class-specific antibodies to<br />

facilitate the analysis <strong>of</strong> mercapturic acids.<br />

Conventional approaches to generate antibodies to low molecular weight<br />

(MW) organic chemicals involves the covalent attachment <strong>of</strong> the small<br />

molecule (hapten) to a strongly antigenic protein, e.g. keyhole limpet


haemocyanin or bovine serum albumin, for the immunisation <strong>of</strong> mice. This<br />

strategy is usually effective in the case <strong>of</strong> strongly antigenic haptens, e.g.<br />

aromatic nitro compounds <strong>of</strong> PCA-DNA adducts. However, all <strong>of</strong> our<br />

attempts to use this approach to generate antibodies against relatively low<br />

MW and weakly antigenic mercapturic acids, e.g. S-(2-hydroxyethyl)-Nacetylcysteine,<br />

failed. Antibodies were generated but these were directed<br />

against the strongly antigenic carrier protein(s).<br />

Covalent binding to macromolecules is believed to provide the basis <strong>of</strong><br />

allergic responses, e.g. skin sensitisation reactions, to small molecules and,<br />

possibly, a basis for the induction <strong>of</strong> auto-immune responses. Thus, the<br />

binding <strong>of</strong> the small molecule transforms normal proteins into ‘foreign’<br />

proteins which trigger an immune response. Recently we have employed<br />

this principle in an attempt to direct the immune response specifically<br />

against mercapturic acid haptens by immunising mice with the haptens<br />

bound to a non-antigenic carrier protein, i.e. mouse serum albumin.<br />

Preliminary results indicate that this tactic has been successful. Overall the<br />

treatment induced fewer antibody-producing cells. However, the antibodies<br />

that were generated show high affinities and specificity toward model<br />

mercapturic acids including S-(2-hydroxyethyl) and S-phenylmercapturic<br />

acid. Studies are in progress to investigate the performance <strong>of</strong> these<br />

antibodies in an immunoenrichment mode.<br />

The preliminary results <strong>of</strong> our studies using non-antigenic protein<br />

carriers are very encouraging and have provided fresh insights which may<br />

assist in directing immune responses against the specific structural features<br />

<strong>of</strong> interest. Improvements in our ability to tailor the antibody will prove<br />

extremely valuable in optimising the properties <strong>of</strong> antibodies to meet<br />

specific needs, e.g. to enrich DNA, protein or mercapturic acid adducts for<br />

application in identifying the chemical initiators <strong>of</strong> human cancer and<br />

quantifying exposures to these agents.<br />

Cancer risk assessment<br />

Human exposure monitoring (determination <strong>of</strong> dose)<br />

The assessment <strong>of</strong> cancer risks posed by exposure to genotoxic chemicals<br />

has two components: determination <strong>of</strong> the dose and determination <strong>of</strong> the<br />

effect (increment in cancer incidence) caused by that dose. The introduction<br />

<strong>of</strong> the target dose concept by Ehrenberg in the early 1970s has provided the<br />

key to modern strategies to assess genotoxic risks (Ehrenberg, 1974, 1979;<br />

Ehrenberg et al., 1974). This new dose concept was developed to provide a<br />

measure <strong>of</strong> the critical dose, i.e. the dose <strong>of</strong> the ultimate genotoxic agent(s)<br />

penetrating to DNA. Target dose is much more relevant to risk assessment<br />

than is exposure dose. The determination <strong>of</strong> target dose automatically

compensates for individual or species differences in the operation <strong>of</strong><br />

metabolic and biokinetic factors that control the quantitative (and<br />

qualitative) relationships between the exposure and the dose <strong>of</strong> the ultimate<br />

toxicant delivered to the target. Measurements <strong>of</strong> target dose may be<br />

applied to improve the extrapolation <strong>of</strong> risk data from experimental<br />

models to humans and may also provide improved definitions <strong>of</strong> risks to<br />

individuals. The determination <strong>of</strong> target dose in humans may be viewed,<br />

therefore, as an approach towards direct risk monitoring as well as a more<br />

relevant approach to monitor human exposures to genotoxic chemicals.<br />

Determination <strong>of</strong> target dose<br />

The determination <strong>of</strong> target dose raises numerous technical and theoretical<br />

problems. Target dose can be determined by measuring primary products,<br />

e.g. DNA adducts, formed when genotoxic agents react with DNA. The<br />

kinetics <strong>of</strong> formation and decay <strong>of</strong> these adducts must also be determined<br />

(vide infra) in order to transform measurements <strong>of</strong> amounts <strong>of</strong> adducts into<br />

estimates <strong>of</strong> target dose. Human tissue DNA is not readily accessible for<br />

monitoring purposes: surrogate dose monitors are required. There are<br />

numerous possibilities including the determination <strong>of</strong> adducts in white<br />

blood cell DNA or <strong>of</strong> the corresponding adducts in the haemoglobin <strong>of</strong><br />

circulating erythrocytes.<br />

Such indirect approaches require validation. Haemoglobin is the most<br />

extensively studied surrogate, not only because <strong>of</strong> its accessibility and<br />

relative abundance but also because <strong>of</strong> the relative stability <strong>of</strong> haemoglobin<br />

adducts and the longevity <strong>of</strong> erythrocytes which permit retrospective<br />

estimates <strong>of</strong> dose received by the erythrocytes over a period <strong>of</strong> about 4<br />

months. Current evidence indicates that all electrophiles that undergo<br />

covalent reactions with DNA also react with haemoglobin. Furthermore<br />

the amounts <strong>of</strong> haemoglobin adducts are quantitatively related to the rates<br />

<strong>of</strong> formation <strong>of</strong> DNA adducts in the tissues. However the proportional<br />

relationships between the doses delivered to tissue DNA and to<br />

haemoglobin or to any other surrogate will vary from chemical to chemical<br />

and will have to be established using experimental models.<br />

Measurement <strong>of</strong> haemoglobin adducts<br />

A.S.WRIGHT ET AL. 189<br />

Genotoxic chemicals undergo covalent reactions with a variety <strong>of</strong><br />

nucleophilic centres in haemoglobin including the sulphydryl group <strong>of</strong><br />

cysteine, the N 1 and N 3 atoms <strong>of</strong> histidine and the amino groups <strong>of</strong> Nterminal<br />

valine residues. Ehrenberg’s group (Osterman-Golkar et al., 1976;<br />

Calleman et al, 1978; Törnqvist et al., 1986a) has pioneered the<br />

development <strong>of</strong> methods to detect, identify and quantify adducts formed at<br />

each <strong>of</strong> these centres. A review <strong>of</strong> these and methods for the analysis <strong>of</strong>


‘labile’ adducts formed, for example, during exposure to aromatic amines<br />

(Green et al., 1984; Albrecht and Neumann, 1985) is beyond the scope <strong>of</strong><br />

this paper (for reviews see Farmer, 1991; Skipper and Naylor, 1991).<br />

However, probably the most powerful and valuable approach was<br />

developed by Törnqvist et al. (1986a, b) who showed that adducts with the<br />

N-terminal valine residues <strong>of</strong> haemoglobin could be specifically enriched by<br />

scission in a modified Edman reaction followed by extraction. This<br />

enrichment procedure greatly facilitates sample analysis by GC/MS.<br />

Immunoassays are also being introduced as alternatives to physicochemical<br />

methods for the determination <strong>of</strong> protein adducts (Wraith et al., 1988).<br />

However, as in the case <strong>of</strong> DNA adducts, the biggest impact <strong>of</strong><br />

immunotechnology on the analysis <strong>of</strong> protein adducts will probably be in<br />

the immunoenrichment <strong>of</strong> low levels <strong>of</strong> adducts for analysis by physicochemical<br />

methods.<br />

Determination <strong>of</strong> biological effects<br />

Tumour incidence<br />

The determination <strong>of</strong> target dose is essential for assessing cancer risks<br />

posed by low-level exposures to genotoxic chemicals. The other requisite is<br />

know ledge <strong>of</strong> the human dose-carcinogenic response relationships in the<br />

low-dose range. The lack <strong>of</strong> intrinsic resolving power <strong>of</strong> classical<br />

epidemiological methods (vide supra) prevents effective applications to<br />

detect small carcinogenic effects associated with low exposures to any<br />

particular genotoxic chemical. Furthermore, the detection limits <strong>of</strong> animal<br />

cancer studies fall short <strong>of</strong> ‘acceptable’ risk limits by three to four orders <strong>of</strong><br />

magnitude (Wright, 1991). This poor sensitivity compels the use <strong>of</strong> high<br />

test doses in order to ensure that significant carcinogens do not go<br />

undetected. However, it is generally accepted that high doses <strong>of</strong> chemicals<br />

may induce tumours by non-specific mechanisms, e.g. via tissue injury and<br />

compensatory cell proliferation, that do not operate at low doses (Ames,<br />

1989; Wright, 1991). Many, if not all genotoxic chemicals induce cell<br />

injury at high (thresholded) doses. Clearly, extrapolation <strong>of</strong> such high dose<br />

risk data to the relevant low-dose range may, at the very least, lead to a<br />

gross overestimation <strong>of</strong> risk.<br />

Determination <strong>of</strong> mutagenic potency<br />

In considering the impact that a low-level exposure to a genotoxic<br />

chemical may have on cancer incidence, it is reasonable to suggest that the<br />

mutagenic propensity <strong>of</strong> the chemical, although <strong>of</strong> a low order, would<br />

nevertheless be the overriding risk factor. Thus, it is probable that any