Toxicology of industrial compounds

Toxicology of Industrial Compounds

Toxicology of Industrial Compounds Edited by

HELMUT THOMAS CIBA-GEIGY Ltd, Basel, Switzerland ROBERT HESS Dornach, Switzerland and FELIX WAECHTER CIBA-GEIGY Ltd, Basel, Switzerland

This edition published in the Taylor & Francis e-Library, 2005. “To purchase your own copy of this or any of Taylor & Francis or Routledge’s collection of thousands of eBooks please go to www.eBookstore.tandf.co.uk.” UK Taylor & Francis Ltd, 4 John Street, London WC1N 2ET USA Taylor & Francis Inc., 1900 Frost Road, Suite 101, Bristol, PA 19007 Copyright © Taylor & Francis Ltd 1995 All rights reserved. No part of this publication may be reproduced, stored in a retrieval system, or transmitted, in any form or by any means, electronic, electro static, magnetic tape, mechanical, photocopying, recording or otherwise, without the prior permission of the copyright owner. Library of Congress Cataloguing Publication data are available Cover design by Hybert Design & Type, Maidenhead, Berks. British Library Cataloguing in Publication Data A catalogue record for this book is available from the British Library. ISBN 0-203-97962-1 Master e-book ISBN

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

Contents

PART ONE

Preface

vii

List of Contributors

ix

Bioavailability and metabolic aspects of industrial chemicals

1

1.

Biomonitoring and Absorption of Industrial Chemicals: the Challenge of Organic Solvents F.A.de Wolff S.Kezic J.G.M.van Engelen A.C.Monster

2

2.

Toxicokinetics and Biodisposition of Industrial Chemicals N.P.E.Vermeulen R.T.H.van Welie B.M.de Rooij J.N.M.Commandeur

12

3.

Metabolic Activation of Industrial Chemicals and Implications for Toxicity G.J.Mulder

36

4.

Sizing Up the Problem of Exposure Extrapolation: New Directions in Allometric Scaling D.B.Campbell

44

PART TWO

Reactive industrial chemicals

59

5.

Metabolism of Reactive Chemicals P.J.van Bladeren B.van Ommen

60

6.

Methods for the Determination of Reactive Compounds P.Sagelsdorff

72

PART THREE Pulmonary toxicology of industrial chemicals 7.

Studies to Assess the Carcinogenic Potential of Man-Made Vitreous Fibers T.W.Hesterberg G.R.Chase R.A.Versen R.Anderson

90 91

v

8.

Pulmonary Toxicity Studies with Man-Made Organic Fibres: Preparation and Comparisons of Size-separated Para-aramid with Chrysotile Asbestos Fibres D.B.Warheit M.A.Hartsky C.J.Butterick S.R.Frame

117

9.

Pulmonary Hyperreactivity to Industrial Pollutants J.Pauluhn

129

10.

Mechanisms of Pulmonary Sensitization I.Kimber

138

11.

Occupational Asthma Induced by Chemical Agents C.A.C.Pickering

149

PART FOUR Biomarkers and risk assessment of industrial chemicals

157

12.

Biomarkers and Risk Assessment K.Hemminki

158

13.

Extrapolation of Toxicity Data and Assessment of Risk N.Fedtke

167

14.

Molecular Approaches to Assess Cancer Risks A.S.Wright J.P.Aston N.J.van Sittert W.P.Watson

180

15.

Evaluation of Toxicity to the Immune System H.-W.Vohr

197

16.

New Strategies: the Use of Long-term Cultures of Hepatocytes in Toxicity Testing and Metabolism Studies of Chemical Products Other than Pharmaceuticals V.Rogiers M.Akrawi S.Coecke Y.Vandenberghe E.Shephard I.Phillips A.Vercruysse

207

PART FIVE 17.

Mechanisms of toxicity of industrial chemicals Peroxisome Proliferation B.G.Lake R.J.Price

222 223

vi

18.

Neurotoxicity Testing of Industrial Compounds: in vivo Markers and Mechanisms of Action K.J.van den Berg J.-B.P.Gramsbergen E.M.G.Hoogendijk J.H.C.M.Lammers W.S.Sloot B.M.Kulig

238

19.

Endocrine Toxicology of the Thyroid for Industrial Compounds C.K.Atterwill S.P.Aylward

255

20.

Testing and Evaluation for Reproductive Toxicity A.K.Palmer

280

PART SIX

Toxicity of selected classes of industrial chemicals

300

21.

Special Points in the Toxicity Assessment of Colorants (Dyes and Pigments) H.M.Bolt

301

22.

Toxicology of Textile Chemicals D.Sedlak

309

23.

Antioxidants and Light Stabilisers: Toxic Effects of 3,5-Dialkyl-hydroxyphenyl Propionic Acid Derivatives in the Rat and their Relevance for Human Safety Evaluation H.Thomas P.Dollenmeier E.Persohn H.Weideli F.Waechter

317

24.

Toxicology of Surfactants: Molecular, Mechanistic and Regulatory Aspects W.Sterzel

339

PART SEVEN Controversial mechanistic and regulatory issues in the safety assessment of industrial chemicals

355

25.

Low Dose of a Genotoxic Carcinogen does not ‘Cause’ Cancer; it Accelerates Spontaneous Carcinogenesis W.K.Lutz

356

26.

Controversial Mechanistic and Regulatory Issues in Safety Assessment of Industrial Chemicals—an Industry Point of View H.-P.Gelbke

362

Index

373

Preface

A large number of chemical compounds are being constantly introduced and produced to ease and comfort modern human life. Among those, the industrial compounds represent that particular fraction of chemicals which are not intended for use in biological systems, but to which humans may be non-intentionally exposed; at the workplace, by product application or through the environment. The International Society for the Study of Xenobiotics (ISSX) committed itself to address, for the first time in the long history of industrial chemicals, the toxicology of this class of compounds in an intensive scientific workshop held June 12 through 15, 1994 in Schluchsee, Germany. This workshop was not only the first such event hosted by ISSX since its foundation in 1981, but also an extension of the society’s scope beyond its traditionally covered objective to promote studies on xenobiotic metabolism, disposition and kinetics mainly of drugs and agrochemicals. The large classes of pharmaceuticals and agrochemicals had been deliberately excluded from the scope of this workshop, since their terms of use generally demand ample registrational toxicity testing that inevitably leads to a wealth of information on, and profound toxicological characterisation of, these compounds. Industrial chemicals, instead, which are frequently produced in large quantities such as pigments, dye-stuffs, plastic materials and additives, detergents, solvents, etc., to name but a few, are in many cases subjected to the examination of a very basic handling safety only, and may lack any further toxicity testing. This implies that essentially nothing is known about their bioavailability, metabolism, excretion and toxicological properties—unless problems arise. And once toxicity problems come up, the question arises with them of whether or not the available and traditionally employed methodology is appropriate to approach and solve them. This, because different from the largely low molecular weight structures developed for use in biological systems, industrial chemicals are often characterised by rather high molecular weight and the incorporation of peculiar structural entities.

viii

Therefore, it was the aim of this workshop to contribute to the investigation of industrial chemicals by focussing on the individual structure, its biological fate, its potential toxicity to mammals and the molecular mechanisms possibly underlying such adverse effects by highlighting the use and significance of experimental toxicology, with special emphasis on mechanistic aspects, in the safety assessment of industrial compounds as well as to current regulatory and legal considerations. Topics had been selected to review generally approved facts and mechanisms, and to particularly address and explore areas of investigative and regulatory uncertainty, thereby intending to bring together the broadly diverse expertise and interests of academic researchers, corporate scientists, experts in safety assessment and representatives from regulatory authorities. The following contributions reflect a substantial selection of the 27 lectures and six short communications presented during the workshop. May they succeed in setting a landmark for the due change from the current era of black-box toxicology and largely undifferentiated regulatory treatment of industrial chemicals to the desirable toxicology and safety assessment by structure in the future. We gratefully acknowledge the substantial financial support by CIBAGEIGY and the RCC Group as well as the financial contributions of ADME Bioanalysis, BASF, Henkel, Hüls, Lonza, Schering and Union Carbide. Our gratitude is also extended to Mrs Ch.Zehnder for secretarial assistance and to Taylor & Francis for continuous support, patience and encouragement to make this publication possible. H.Thomas R.Hess F.Waechter

Contributors

May Akrawi Department of Biochemistry and Molecular Biology, University College London, Gower Street, London WC1E 6BT, UK Robert Anderson Schuller MTC, Health, Safety and Environmental Department, Toxicology Group, PO Box 625005, Littleton, CO 80162–5005, USA J.Paul Aston Sittingbourne Research Centre, Sittingbourne, Kent ME9 8AG, UK Christopher K.Atterwill CellTox Centre, University of Hertfordshire, Hatfield Campus, College Lane, Hatfield AL10 9AB, UK Samuel P.Aylward CellTox Centre, University of Hertfordshire, Hatfield Campus, College Lane, Hatfield AL10 9AB, UK Peter J.van Bladeren TNO Nutrition and Food Research, PO Box 360, Utrechtseweg 48, NL-3700 AJ Zeist, The Netherlands Hermann M.Bolt Institut für Arbeitsphysiologie, Universität Dortmund, Ardeystrasse 67, D-44139 Dortmund, Germany Charles J.Butterick Texas Technical Health Sciences Centre, Lubbock, TX, USA D.Bruce Campbell Servier Research and Development, Fulmer Hall, Windmill Road, Fulmer, Slough SL3 6HH, UK Gerald R.Chase Schuller MTC, Health, Safety and Environmental Department, Toxicology Group, PO Box 625005, Littleton, CO 80162–5005, USA

x

Sandra Coecke Vrije Universiteit Brussel, Department of Toxicology, Laarbeeklaan 103 B-1090, Brussels, Belgium Jan N.M.Commandeur Division of Molecular Toxicology, Department of Pharmacochemistry, Vrije Universiteit van Amsterdam, De Boelelaan 1083, H1081 NL-V Amsterdam, The Netherlands Peter Dollenmeier CIBA-GEIGY Ltd., R-1002.2.62, PO Box CH-4002 Basel, Switzerland Jacqueline G.M.van Engelen Coronel Laboratory, University of Amsterdam, Academic Medical Centre, Meibergdreef 15, NL-1105 Amsterdam, The Netherlands Norbert Fedtke Hüls AG, Bau 2328/PB 12, D-45764 Marl, Germany Steven R.Frame DuPont Central Research and Development, Haskell Laboratory, PO Box 50, Elkton Road, Newark, DE 19714–0050, USA Heinz-Peter Gelbke BASF AG, Abt. Toxikologie, D-67056 Ludwigshafen, Germany Jan-Bert P.Gramsbergen Department of Public Health, Erasmus University, Rotterdam, The Netherlands Mark A.Hartsky DuPont Central Research and Development, Haskell Laboratory, PO Box 50, Elkton Road, Newark, DE 19714–0050, USA Kari Hemminki CNT, Karolinska Institute, Novum, S-141 57 Huddinge, Sweden Thomas W.Hesterberg Schuller MTC, Health, Safety and Environmental Department, Toxicology Group, PO Box 625005, Littleton, CO 80162–5005, USA Elisabeth M.G.Hoogendijk TNO Toxicology, Department of Neurotoxicology, PO Box 5815, NL-2280 HV Rijswijk, The Netherlands Sanja Keži Coronel Laboratory, University of Amsterdam, Academic Medical Centre, Meibergdreef 15, NL-1105 Amsterdam, The Netherlands Ian Kimber Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire SK10 4TJ, UK

xi

Beverly M.Kulig TNO Toxicology, Department of Neurotoxicology, PO Box 5815, NL-2280 HV Rijswijk, The Netherlands Brian G.Lake BIBRA International, Woodmansterne Road, Carshalton, Surrey, SM5 4DS, UK Jan H.C.M.Lammers TNO Toxicology, Department of Neurotoxicology, PO Box 5815, NL-2280 HV Rijswijk, The Netherlands Werner K.Lutz Universität Würzburg, Institut für Toxikologie, Versbacher Strasse 9, D-97078 Würzburg, Germany Aart C.Monster Coronel Laboratory, University of Amsterdam, Academic Medical Centre, Meibergdreef 15, NL-1105 Amsterdam, The Netherlands Gerard J.Mulder Center for Bio-Pharmaceutical Sciences, Sylvius Laboratories, Leiden University, PO Box 9503, NL-2300 RA Leiden, The Netherlands Ben van Ommen TNO Nutrition and Food Research, PO Box 360, Utrechtseweg 48, NL-3700 AJ Zeist, The Netherlands Anthony K.Palmer Huntingdon Research Centre Ltd., PO Box 2, Huntingdon, Cambs, PE18 6ES UK Jürgen Pauluhn BAYER AG, Department of Toxicology, Institute of Industrial Toxicology, Bldg. 514, D-42096 Wuppertal, Germany Elke Persohn CIBA-GEIGY Ltd., Cell Biology Unit, R-1058.2.64, PO Box, CH-4002 Basel, Switzerland Ian Phillips Department of Biochemistry, Queen Mary and Westfield College, University of London, Mile End Road, London, E1 4NS, UK C.A.C.Pickering North West Lung Centre, Wythenshawe Hospital, Southmoor Road, Manchester M23 9LT, UK Roger J.Price BIBRA International, Woodmansterne Road, Carshalton, Surrey SM5 4DS, UK

xii

Vera Rogiers Vrije Universiteit Brussel, Department of Toxicology, Laarbeeklaan 103, B-1090 Brussels, Belgium Ben M.de Rooij Division of Molecular Toxicology, Department of Pharmacochemistry, Vrije Universiteit van Amsterdam, De Boelelaan 1083, H1081 NL-V Amsterdam, The Netherlands Peter Sagelsdorff CIBA-GEIGY Ltd., Cell Biology Unit, R-1058.2.52, PO Box, CH-4002 Basel, Switzerland Dieter Sedlak Enviro Tex GmbH, Provinostrasse 52, D-86153 Augsburg, Germany Elizabeth Shephard Department of Biochemistry and Molecular Biology, University College London, Gower Street, London WC1E 6BT, UK Nico J.van Sittert Sittingbourne Research Centre, Sittingbourne, Kent ME9 8AG, UK Willem S.Sloot TNO Toxicology, Department of Neurotoxicology, PO Box 5815, NL-2280 HV Rijswijk, The Netherlands Walter Sterzel Henkel KGaA, TTB-Toxikologie, Geb. Z33, D-40191 Düsseldorf, Germany Helmut Thomas CIBA-GEIGY Ltd., Cell Biology Unit, R-1058.2.46, PO Box, CH-4002 Basel, Switzerland. Current address: Ciba-Pharmaceuticals, Stamford Lodge, Wilmslow, Cheshire SK9 4LY, UK Kornelis J.van den Berg TNO Toxicology, Department of Neurotoxicology, PO Box 5815, NL-2280 HV Rijswijk, The Netherlands Yves Vandenberghe Vrije Universiteit Brussel, Department of Toxicology, Laarbeeklaan 103 B-1090 Brussels, Belgium Antoine Vercruysse Vrije Universiteit Brussel, Department of Toxicology, Laarbeeklaan 103 B-1090 Brussels, Belgium Nico P.E.Vermeulen Division of Molecular Toxicology, Department of Pharmacochemistry, Vrije Universiteit Amsterdam, De Boelelaan 1083, H1081 NL-V Amsterdam, The Netherlands

xiii

Richard A.Versen Schuller MTC, Health, Safety and Environmental Department, Toxicology Group, P.O. Box 625005, Littleton, CO 80162–5005, USA Hans-Werner Vohr Bayer AG, Fachbereich Toxikologie, Institut für Toxikologie Landwirtschaft, Friedrich-Ebert-Strasse 217, D-42096 Wuppertal, Germany Felix Waechter CIBA-GEIGY Ltd, Cell Biology Unit, R-1058.2.68, PO Box, CH-4002 Basel, Switzerland David B.Wahrheit DuPont Central Research and Development, Haskell Laboratory, PO Box 50, Elkton Road, Newark, Delaware 19714–0050, USA William P.Watson Sittingbourne Research Centre, Sittingbourne, Kent ME9 8AG, UK Hansjörg Weideli CIBA-GEIGY Ltd, R-1002.2.59, PO Box, CH-4002 Basel, Switzerland Ronald T.H.van Welie Division of Molecular Toxicology, Department of Pharmacochemistry, Vrije Universiteit van Amsterdam, De Boelelaan 1083, H1081 NL-V Amsterdam, The Netherlands Frederik A.de Wolff Coronel Laboratory, University of Amsterdam, Academic Medical Centre, Meibergdreef 15, 1105 Amsterdam, The Netherlands Alan S.Wright Sittingbourne Research Centre, Sittingbourne, Kent ME9 8AG, UK

PART ONE Bioavailability and metabolic aspects of industrial chemicals

1 Biomonitoring and Absorption of Industrial Chemicals: the Challenge of Organic Solvents FREDERIK A.DE WOLFF*, SANJA KEŽI , JACQUELINE G.M.van ENGELEN and AART C.MONSTER University of Amsterdam, Academic Medical Center, Amsterdam Introduction Organic solvents form a very important group of industrial chemicals. They are widely used in a range of occupational settings and may exert a number of deleterious effects when subjects are acutely or chronically exposed. Among the acute effects are skin and mucosal irritation and general anaesthesia produced by most solvents at high air concentrations. Examples of chronic effects are peripheral neuropathy after long-term exposure to n-hexane or carbon disulphide, and the organo-psychosyndrome or ‘solvent dementia’ which may occur after chronic occupational exposure to a variety of volatile organic compounds. In order to prevent workers from developing solvent-induced occupational disease, it is essential to set standards for the duration and the level of external exposure. For a scientifically based standard, a clear understanding is required of the relationship between external exposure, the uptake by the body, the metabolic fate and the internal dose of the substance. The purpose of this contribution is to demonstrate the value of biokinetic studies in humans to provide a sound scientific basis for regulatory decisions on occupational standards. Biological monitoring In occupational health practice, monitoring is a tool to protect workers from developing chemically-induced disease. Monitoring in preventive health care is described as ‘a systemic continuous or repetitive healthrelated activity, designed to lead if necessary to corrective action’. In occupational health, a complete monitoring programme consists of four parts: environmental, biological and biological effect monitoring, and

* Also: University Hospital of Leiden, Leiden. The Netherlands

F.A.DE WOLFF ET AL. 3

health surveillance. The latter is a major task for the occupational health physician, but biological monitoring and biological effect monitoring are fields of interest to the occupational toxicologist. In this contribution, only biological monitoring will be expounded upon. Biological monitoring (BM) is defined as the ‘measurement and assessment of workplace agents or their metabolites either in tissues, secreta, excreta or any combination of these to evaluate exposure and health risk compared to an appropriate reference’ (Zielhuis & Henderson, 1986). This means that a biological monitoring programme is not limited to the assay of xenobiotics in biological samples. As in clinical laboratory medicine, the pre-analytical phase of the process is very important, and even more so the post-analytical phase of the laboratory analysis, which means the interpretation of the analytical data in biomedical terms. The ultimate goal of biological monitoring is the evaluation of the health risk of workers by estimation of the internal dose of a chemical. This is not limited to measurement of the quantity of the substance absorbed by the body, but may also include the assay of metabolites of toxicological interest, if possible in or near a critical organ (Monster & van Hemmen, 1988). This implies that the absorption, metabolism and elimination of a substance in man should be known before a biological monitoring programme can be performed in practice. Animal experiments are of limited value; volunteer studies in order to determine pulmonary and dermal uptake of organic solvents provide more relevant data for this purpose. Owing to the existence of very sensitive analytical methods it is possible to study the kinetics and metabolism of solvents in volunteers who are experimentally exposed to levels at or far below the official threshold limit values, so that any health risk for the volunteers can almost totally be excluded. As with biological monitoring of most other substances, in the case of organic solvents the compound itself and/or its metabolite in blood or urine can be measured. Studies with volatile, rather lipophilic, substances have an additional advantage, namely that the solvent can also be measured in expired air. Analytically this has the advantage of an extremely clean matrix in comparison with body fluids, whereas biologically, air samples provide us with information on the blood concentration of a volatile compound. Moreover, collection of expired air is non-invasive and large volumes are readily available (Droz & Guillemin, 1986). An example of a study on solvents in volunteers is the one carried out in our laboratory on the biokinetics of n-hexane and its neurotoxic metabolite 2,5-hexanedione (Van Engelen et al., in preparation). Volunteers are exposed during 15 min to 60 ppm hexane by inhalation. The minute volume and the respiratory rate are measured and blood and exhaled air sampled frequently for determination of 2,5-hexanedione and n-hexane,

4 BIOMONITORING AND ABSORPTION OF INDUSTRIAL CHEMICALS

respectively. Each volunteer is exposed twice in succession on one test day in order to get an impression of the within-day intra-individual variation. Venous blood is sampled through a catheter, and alveolar air is collected after holding breath for 30 s (to achieve equilibrium between pulmonary blood and air) by exhaling through a glass tube which is stoppered immediately. These tubes contain 70 ml alveolar air and the total volume is analyzed for n-hexane by using a purge-and-trap system. 2,5-Hexanedione in serum is measured by using electron capture detection after derivatization, with a detection limit of 30 micro-mol l−1 (Keži and Monster, 1991). During exposure the concentration of n-hexane in alveolar air increases very rapidly and decreases after discontinuation of exposure. The half-life time of exhalatory elimination after the distribution phase is in the order of 30 min. 2,5-Hexanedione becomes detectable in blood as fast as 2–3 min after commencement of n-hexane exposure. After discontinuation of dosing the metabolite concentration continues to increase for another 3 min, to disappear from the plasma with a half-life of approximately 1.5 h. The second exposure period on the same day shows very reproducible n-hexane and 2,5-hexanedione curves in the same individual. Between individuals there is considerable variation in kinetics and metabolism, and this issue is being studied in detail at present. Before a biological monitoring programme can be designed, a detailed biokinetic study like this one, of every solvent being used in industry, has to be performed. Without kinetic data it is impossible to choose for instance the correct matrix, the compound to be measured, or the sampling frequency and time. In addition, these data are necessary to establish a relationship between ambient air concentrations of a chemical (external exposure), and the biological parameters used to estimate a health risk. Absorption The primary association of the pharmacologist or general toxicologist, when reading or hearing the term ‘absorption’, is with ‘intestinal’. For drugs, gastrointestinal uptake is indeed the most common route to enter the body. In case of occupational exposure, however, intestinal absorption is of minor importance. The occupational toxicologist is, therefore, more inclined to pay attention to entry routes other than the intestine, the most important being pulmonary and dermal uptake. Pulmonary uptake There are a number of parameters which affect the pulmonary uptake of organic solvents. In the first place, the physical chemistry of the compound is of importance. Both the blood-to-gas and the tissue-to-blood partition

F.A.DE WOLFF ET AL. 5

Figure 1.1 The mean minute volume (1 min−1) and the percentage of the minute volume cleared from solvent (shaded area) during exposure to styrene (left) and 1,1, 1-trichloroethane (right) at increasing degree of workload.

coefficients determine the absorption through the alveolar membrane and the distribution over the body. Furthermore, exercise is an important physiological determinant. With increasing exercise, ventilation increases and, therefore, also the availability of the vapour to the lung per unit of time. In addition, cardiac output increases during exercise, and this may affect absorption, distribution and metabolism through enhanced blood flow. Finally, the elimination of a solvent which occurs during exposure may significantly affect the uptake rate. The percentage of the vapour not retained by the body but exhaled again is dependent on, again, physicochemical factors such as solubility, but also on the rate of metabolism (Fiserova-Bergerova, 1985). In order to demonstrate the different factors which may affect pulmonary absorption of vapours we have constructed Figure 1.1, based on earlier work of Astrand et al. (Astrand, 1975). In their studies, volunteers were exposed to different vapours such as styrene or 1,1,1trichloroethane at increasing degrees of workload during 2 h. The first 30 min they were exposed at rest, and then the workload was increased every 30 min with 50 W. The minute volume, here referred to as ‘supply’, was measured and expressed in 1 min−1, and the exhaled solvent concentration was also measured at regular intervals. The shaded area of the vertical bars in Figure 1.1 indicate the percentage of minute volume cleared from the solvent, averaged over the observation period. This is considered to be a measure for pulmonary uptake.

6 BIOMONITORING AND ABSORPTION OF INDUSTRIAL CHEMICALS

During continuous exposure to a constant concentration and at increasing exercise the uptake of styrene remains constant, expressed in terms of percentage of the minute volume cleared. Apparently, the body is not easily saturated with styrene. The picture for 1,1,1-trichloroethane is completely different. Although the minute volume at each level of workload is comparable with that of the styrene experiment, it is clear that the retention of 1,1,1-trichloro-ethane is much lower. Apparently, the body becomes rapidly saturated with 1,1,1-trichloroethane. The reasons for the difference in pulmonary uptake between these two solvents are evident. Styrene is highly soluble in blood and it is extensively metabolized to mandelic acid and phenyl glyoxylic acid. The retention in the body remains the same, and therefore the uptake increases proportionally with the minute volume. In contrast to styrene, 1,1,1-trichloroethane has only a limited solubility in blood, and it is hardly metabolized. This means that during exposure the body becomes rapidly saturated with the substance, and that an increase in minute volume by increasing workload results in a lower retention, and hardly in higher uptake. Differences in kinetic behaviour, as demonstrated for styrene and 1,1,1-trichloroethane, are important for the design of a biological monitoring programme. Dermal uptake Absorption of solvents through the skin may be affected by a number of factors. Many organic solvents are able to penetrate the skin and thus enter the body. This is a rather well-known fact which can be prevented in industrial practice by use of protective clothing. It is, however, less common knowledge that solvents in the vapour phase may also penetrate the skin. In case of skin exposure to liquids usually a small surface is exposed, whereas in case of vapour the whole body surface of about 2 m2 may be exposed. This means that under certain conditions skin absorption of vapour may significantly contribute to the amount absorbed by inhalation. Other parameters which may affect skin absorption are the temperature, and the ability of some solvents to increase their own absorption by causing skin hyperaemia through irritation. To demonstrate these factors, some preliminary results are shown of a volunteer study on skin penetration of solvents in the liquid and vapour phases (Keži et al., in preparation). The experimental conditions are as follows. The volunteer is seated in a clear-air cabin in order to avoid additional inhalatory exposure to vapour in the experimental room. The arm is the only part of the body outside the cabin. In case of exposure to liquid on the skin, the solvent is put in a chamber which is pressed on to the skin during the exposure period, which

F.A.DE WOLFF ET AL. 7

is usually no longer than a few minutes. The exposed area is usually in the order of 20 cm2. In the case of dermal exposure to vapour, the volunteer places the lower arm into a piece of drainage pipe through which the vapour is led with controlled flow and concentration in air. Uptake of liquid or vapour is measured in both cases by determination of the solvent in expired air, by the sampling method described earlier. Figure 1.2 shows the dermal uptake and elimination of two different liquids in one volunteer. A surface of 27 cm2 was exposed during 3 min to pure 1,1.1-trichloroethane and to tetrachloroethene. It is clear that 1,1,1trichloroethane is absorbed through the skin much faster and to a much greater extent than tetrachloroethene, at least in exposure to the liquids. However, when the skin is exposed to the same solvents in the vapour phase the picture becomes totally different. Here the lower arm, which has a surface of about 500 cm2, was exposed during 15 min to solvent concentrations of approximately 500 µmol 1−1 air (Figure 1.3). In the case of vapour exposure no difference in absorption kinetics is observed, and only a small difference in expired air concentration is seen. The reason for the discrepancy between vapour exposure is that 1,1,1trichloro-ethane causes skin irritation as the liquid, but not in the vapour phase. Irritation leads to hyperaemia and, hence, increased absorption. As it is known that dermal exposure to vapour may lead to detectable absorption, the contribution of vapour uptake of the skin in comparison to inhalatory absorption should be evaluated. This was done with trichloroethene as an example (Figure 1.4). Both curves were obtained in the same volunteer. The dermal exposure was performed first, followed by the inhalatory test after a wash out period of 2 weeks. The exposure period was 15 min, and the inhalatory concentration was 4.1 µmol l−1. Dermal exposure of the lower arm took place at 1.4 mmol l−1. It appears that uptake from the lungs occurs much faster than via the skin. This is conceivable because the stratum corneum is a stronger barrier than the alveolar epithelium, and causes a shift to the right of the tmax. It can also be seen that inhalatory exposure leads to a much higher expired air concentration than dermal exposure. But in this respect we should realize that only a small part of the skin was exposed, namely about 500 cm2. In fact the result should be extrapolated to the total surface of the human skin, which is about 2 m2. These results indicate that dermal exposure to solvent vapour should not be neglected when the safety of the industrial environment is evaluated. This is of special importance when ambient air concentrations are high, and workers are protected with protective masks but not with gloves. Another example in which skin absorption may be high in comparison with inhalation are those solvents which are readily absorbed by the skin, such as 2-butoxyethanol (Johanson and Boman, 1991).

8 BIOMONITORING AND ABSORPTION OF INDUSTRIAL CHEMICALS

Figure 1.2 Elimination of 1,1,1-trichloroethane and tetrachloroethene by expired air after dermal exposure to the liquid of 27 cm2 fore-arm skin during 3 min. 1,1,1trichloroethane liquid irritates the skin.

Figure 1.3 Elimination of 1,1,1-trichloroethane and tetrachloroethene by expired air after dermal exposure to the vapour of 500 cm2 lower-arm skin during 15 min to 500 µmol l−1 air.

The temperature of the solvent is another factor that may have an influence on uptake through the skin. Figure 1.5 shows the results of dermal exposure to liquid tetrachloroethene and n-hexane at two different temperatures in one volunteer. Exposure time here was only 1 min, and absorption and elimination were measured by analysis of the vapours in expired air.

F.A.DE WOLFF ET AL. 9

Figure 1.4 Elimination of trichloroethene by expired air during and after inhalatory exposure to 4.1 µmol l−1 trichloroethene during 15 min, and after dermal vapour exposure during 15 min of the lower-arm skin (500 cm2 to 1.4 mmol l−l air).

At the low temperature of the liquid (15°C), the uptake of tetrachloroethene is negligible when compared with a normal skin temperature of 33°C. In case of n-hexane, under comparable circumstances and in the same volunteer, the effect of temperature is much less pronounced. Apparently, the physicochemical properties of the solvent are an additional determining factor. The mechanism on which the difference between tetrachloroethene and n-hexane is based is the subject of further study. Conclusions In occupational health practice, the major absorption routes for organic solvents are not ingestion, but inhalation and skin penetration, the latter both as liquid and as vapour. The physical chemistry of the compound, exercise, and the elimination rate may affect pulmonary uptake. Factors affecting dermal uptake are the ability of the solvent to penetrate the skin as liquid or vapour, the temperature of the liquid, and the irritability of the chemical to the skin. Before a biological monitoring programme for solvent exposure can be set up, the kinetics and metabolism of the various solvents in man should be known. Owing to the availability of sensitive analytical methods it is usually possible to perform volunteer studies at safe exposure levels. Measurement of solvents in expired air and of their metabolites in body fluids is of the utmost importance to estimate the internal dose of the solvents and health risk to which man can be exposed in the work and general environment.

10 BIOMONITORING AND ABSORPTION OF INDUSTRIAL CHEMICALS

Figure 1.5 Elimination of tetrachloroethene and n-hexane by expired air after dermal exposure during 1 min to liquid at 15°C and 33°C

References ǺSTRAND, I., 1975, Uptake of solvents in the blood and tissues in man. A review, Scand J Work Environ Health, 1, 199–218. DROZ, P.O. and GUILLEMIN, M.P., 1986, Occupational exposure monitoring using breath analysis, J Occup Med, 28, 593–602.

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FISEROVA-BERGEROVA, V., 1985, Toxicokinetics of organic solvents, Scand J Work Environ Health, 11, suppl. 1, 7–21. JOHANSON, G. and BOMAN, A., 1991, Percutaneous absorption of 2butoxyethanol vapour in human subjects, Br J Ind Med, 48, 788–92. KEŽI , S. and MONSTER, A.C., 1991, Determination of 2,5-hexanedione in urine and serum by gaschromatography after derivatization with O(pentafluorobenzyl)-hydroxylamine and solid-phase extraction, J Chromatogr, 563, 199–204. MONSTER, A.C. and VAN HEMMEN, J.J., 1988, Screening models in occupational health practice of assessment of individual exposure and health risk by means of biological monitoring in exposure to solvents, In Notten, W.R.F., Herber, R.F. M., Hunter, W.J. et al. (Eds) Health Surveillance of lndividual Workers Exposed to Chemical Agents, pp. 47–53, Berlin: Springer. ZIELHUIS, R.L. and HENDERSON, P.Th., 1986, Definitions of monitoring activities and their relevance for the practice of occupational health, Int Arch Occup Environ Health, 57, 249–57.

2 Toxicokinetics and Biodisposition of Industrial Chemicals NICO P.E.VERMEULEN, RONALD T.H.van WELIE, BEN M.de ROOIJ and JAN N.M.COMMANDEUR Vrije Universiteit, Amsterdam

Introduction In our industrialized world with increasing numbers of body foreign chemicals (xenobiotics) including drugs, food additives, pesticides, industrial chemicals and environmental pollutants, public concern about possible adverse (health) effects is growing. In 1989, for example, actual environmental topics in the Netherlands were photochemical summersmog and the presence of dioxines in milk of cows feeding in the neighbourhood of household refuse combustion furnaces and cable stills (CCRX, 1989). In this regard, most attention is paid to exposure to potentially mutagenic and carcinogenic xenobiotic chemicals. Apart from environmental exposure, especially at the workplace man may be exposed to elevated levels of mixtures of known or unknown chemicals. Two centuries ago, cancer of the scrotum and testicles in chimney-sweepers was the first recognized occupational cancer (Pott, 1795). Since then numerous other hazardous occupational activities have been traced (Farmer et al., 1987). Nowadays, toxicologists are more and more focussed on the in vivo and in vitro bioactivation and bioinactivation mechanisms of chemicals. In the development of toxicity different stages are generally being distinguished: (1) toxicokinetics (absorption, distribution and elimination), (2) biotransformation, resulting in activation or inactivation of the chemicals, (3) reversible or irreversible interactions with cellular or tissue components, (4) protection and repair mechanisms and (5) nature and extent of the toxic effect for the organism (Vermeulen et al., 1990). Knowledge of for example species, dose, route of absorption, time of exposure, tissue and organ selective interactions with (critical) cellular macro-molecules contributes to the understanding of molecular mechanisms of toxicity. Molecular mechanisms are useful in the prediction and prevention of chemically induced toxicities and they may play an important role in for example risk assessment and in the development of safer chemicals (Vermeulen et al., 1990).

N.P.E.VERMEULEN ET AL. 13

In this chapter, first the basic toxicokinetic concepts concerning the dis tribution, elimination and biotransformation of xenobiotics will be summarized. Subsequently, the relevance of these concepts will be illustrated and evaluated with the aid of a number of toxicokinetic studies in animals and humans concerning the nematocide 1,3-dichloropropene, the fungicide etridiazol, the chemical monomer 1,3-butadiene and the industrial solvent, 1,1,2-tri-chloroethylene. Apart from interspecies differences in the toxicokinetics, special attention will be given to interindividual differences in the toxicokinetics, among other things, as a result of genetically determined deficiencies in biotransformation enzymes as well as to its importance for the risk assessment of human exposure to industrial chemicals. Disposition of xenobiotics The overall fate of xenobiotics in an organism is determined by various toxicokinetic processes notably the route of administration, absorption, distribution and elimination. Chemicals may enter the body via various routes. Main routes are the lung, skin and gastrointestinal tract. The intraperitoneal, intramuscular, intravenous and subcutaneous routes are largely confined to experimental toxicological and therapeutic agents. Following absorption, xenobiotics enter the systemic or portal blood circulation. Distribution of chemicals in blood, organs and tissues usually occurs rapidly. The final plasma concentration depends on the ability of the chemicals to pass cell membranes and on their affinity to various macromolecular proteins and tissues. Distribution to the kidney may result in direct excretion of the unchanged parent chemical. The physicochemical characteristics, such as lipophilicity and binding to plasma proteins, play an important role in the ultimate fate of a chemical in the body. The disposition of xenobiotics in the body is shown schematically in Figure 2.1. Its schematic relationship with biological/ toxicological effects is shown in Figure 2.2. Biotransformation plays an important role in the disposition of xenobiotics in vivo. The liver is quantitatively the most important organ in the process of biotransformation. It receives a relative high bloodflow directly from the gastrointestinal tract via the portal vein, sometimes giving rise to the so-called hepatic ‘first-pass effect’ due to the presence of high concentrations of phase I and phase II metabolizing enzymes. Other important organs in biotransformation are the lungs, kidneys and the intestine. The primary object of biotransformation generally is to increase the hydrophilicity of chemicals, thus facilitating excretion by the kidneys in the urine or by the liver in the bile. Phase I reactions involve oxidation, reduction and hydrolysis reactions and phase II reactions conjugation or synthetic reactions. Phase I metabolic reactions generally

14 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS

Figure 2.1 Schematic representation of the fate of xenobiotics in the body according to their physico-chemical properties. Phase I and phase II represent the biotransformation processes. Adapted from Ariens and Simonis (1980).

convert xenobiotic chemicals to more hydrophilic derivatives by introducing functional groups such as hydroxyl, sulphydryl and amino- or carboxylic acid groups. Phase II reactions are conjugation reactions in which the parent compounds or phase I derived metabolites are covalently bound to for example glucuronic acid, sulphate or glutathione. The group of cytochrome P-450 isoenzymes is the most important enzyme system in the catalysis of phase I reactions. The microsomal cytochrome P-450 system consists of various cytochrome P-450 isoenzymes and NADPH-cytochrome P450 reductase. It is involved in different metabolic reactions. At least three main types of activities can be distinguished, namely monooxygenase activity, oxidase activity and reductive activity (Guengerich 1994; Koymans et al., 1993). Glucuronic acid conjugation, catalyzed by UDP-glucuronyltransferases, represents one of the major phase II conjugation reactions in the conversion of exogenous and

N.P.E.VERMEULEN ET AL. 15

Figure 2.2 Disposition and biological effects of xenobiotics subdivided into three phases.

endogenous chemicals. In mammals, another important conjugation reaction of hydroxyl groups is sulfatation, catalyzed by sulfotransferases (Sipes and Gandolfi, 1986). The group of glutathione S-transferase (GST) isoenzymes also represents an important phase II enzyme system. GST isoenzymes consist of two subunits on which the nomenclature is based (Warholm et al., 1986). The most important activity of GSTs is the catalysis of the conjugation of electrophilic, hydrophobic chemicals with the tripeptide glutathione (GSH). In general, GSH conjugation ultimately leads to the urinary excretion of mercapturic acids (N-acetyl-L-cysteine Sconjugates) (Vermeulen, 1989; Van Welie et al., 1992). Toxicokinetic principles General principles The time course for the absorption, distribution, metabolism and elimination of a toxic substance is the subject of toxicokinetics. Implicit in any toxicokinetic description is the assumption that the response of target tissues or organs can be related to concentration profiles of the active form of the substance in that tissue or organ. Furthermore, it is often assumed that blood or plasma concentrations in one way or the other will reflect target tissue or organ concentrations and by inference the toxic effects. Under normal conditions one is generally dealing with first-order or linear kinetics, meaning that the amount of compound absorbed or eliminated (dQ) per unit of time (dt) is proportional to the total amount of compound present in the body. Zeroorder or non-linear kinetics may be valid as a consequence of various causes, e.g. saturation of binding of the toxic substance to plasma proteins or tissue components, or, more frequently

16 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS

Table 2.1 Frequently used toxicokinetic parameters and their formulas

occurring, saturation of biotransformation enzyme systems. For the (mathematical) description of the toxicokinetics of substances, there exist at least two approaches at the moment: the traditional compartment pharmaco(toxico-)kinetic approach, in which the body is divided into one or more compartments, which do not necessarily correspond to physiological or anatomical units, and the physiologically-based pharmaco (toxico-)kinetic approach (PBPK or PBTK), in which organs, tissues and blood flow are taken into consideration. In Table 2.1 a summary of the most important and most frequently used traditional toxicokinetic parameters is shown. The value of some of these parameters is illustrated below, with the examples of 1,3-dichloropropene and etridiazol. The PBPK/PBTK approach is illustrated with the example of 1,3-butadiene. Principles of urinary excretion Of special interest in relation to this contribution also is the urinary excretion of xenobiotics and their metabolites by the kidneys. Two basic

N.P.E.VERMEULEN ET AL. 17

processes, namely glomerular filtration and tubular secretion are used by the kidneys to remove chemicals from the bloodstream into the urine (Hook and Hewitt 1986). The kidneys are highly vulnerable to potential toxicants not only because they receive a high bloodflow (25% of the cardiac output), but also because they have the intrinsic ability to concentrate compounds. Recently, it has also become clear that xenobiotics may become nephrotoxic in the kidney itself due to bioactivation processes in combination with insufficient protection mechanisms (Commandeur and Vermeulen, 1991). The elimination of chemicals by the kidney is generally governed by firstorder processes. During first-order excretion kinetics the urinary elimination rate of a chemical is directly proportional to the plasma concentration. This means that the higher the plasma concentration the more of the chemical will be excreted in urine per unit of time. The urinary elimination rate (dQ/dt) can be calculated from a semi-logarithmic plot of the urinary elimination rate versus the time of the intermittently collected urine samples (dQ/dt (mg h−l)=volume (1)×concentration (mg 1−1)/time (h)) (Figure 2.3A). From the slope of the semi-logarithmic plasma concentration or urinary excretion rate versus time curve, the elimination rate constant (kel) and the urinary half-life of elimination (t1/2) can be calculated. The half-life of elimination is the time required to decrease the plasma concentration or the urinary elimination rate by one-half. The volume of distribution of the chemical normally can not be calculated from the urinary excretion data. Because the amount of chemical excreted in urine per unit of time (dQ/dt) is proportional to the plasma concentration (Cp), the t1/2 derived from the urinary elimination rate constant is identical to the t1/2 of the chemical in plasma. It is evident that under these conditions the urinary excretion rate curve has the same shape as the plasma concentration curve (Figure 2.3B). In practice, the concentration of a chemical in urine (mg l−1) can be determined and multiplied by the volume (1) of the urine sample in order to calcu late the amount (mg) of chemical excreted over a period of time. In a semi-logarithmic plot the amount of chemical excreted is plotted against the midpoint of the interval of collection (Figure 2.3B). The accuracy of the method strongly depends on the way and the number of urine samples collected. As a rule of thumb, urine samples have to be collected during at least four half-lives of elimination. The complete cumulative urinary excretion of a chemical can be calculated as the area under the urinary excretion rate versus time curve including extrapolation time to infinity. Occupational exposure to chemicals frequently occurs 5 days a week, 8 h a day, with an exposure free period of 16 h. Intermittent exposure to a chemical may lead to different accumulation situations in the body depending on the periods between exposure in relation to t1/2 (Table 2.1). No accumulation will occur when the intervals between the exposure

18 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS

Figure 2.3 Schematic representation of first order kinetics of (A) the plasma concentration (Cp) of a chemical versus the urinary elimination rate (dQ/dt), (B) the relation between the elimination rate in plasma and urine and (C) the cumulative excretion ( (%)) versus time. In (B): slope=–kel/2.303 and t1/2=0.693/kel.

Figure 2.4 Urinary excretion of a hypothetical metabolite during 3 days of intermittent exposure: t1/2<16 h. AUC denotes area under the curve. The complete cumulative excretion on day 2 can be calculated as: AUC-2total=AUC-2+AUC-2 −AUC-1.

periods are larger than t1/2. However, accumulation will occur until a steady state is reached when the exposure free interval is shorter than or equal to t1/2. Of interest within the context of this contribution is the first situation in which no accumulation occurs. In occupational practice a situation as presented in Figure 2.4 may often occur. Daily intermittent exposure periods do not lead to accumulation, but the urinary excretion of a xenobiotic or a metabolite is not completed within the exposure free period. Urinary excretion due to the exposure on day 1 continues on day 2, exposure on day 2 continues on day 3, etc. Being interested in the exposure of day 2, the contribution of urinary excretion of day 1 to that of day 2 will have to be subtracted whereas the excretion on day 3 will have to be added. Knowing an individual’s elimination rate constant (kel) and the

N.P.E.VERMEULEN ET AL. 19

urinary concentration at certain time points the net total cumulative excretion of day 2 can be calculated. Monitoring in occupational toxicology In occupational toxicology generally four monitoring approaches are distinguished, namely: environmental monitoring (EM), biological monitoring (BM), biological effect monitoring (BEM) and health surveillance (HS) (Figure 2.5). EM and BM are concerned with the measurement and assessment of ambient exposure and health risk compared to appropriate references. EM determines xenobiotics at the workplace, BM determines xenobiotics or their metabolites in tissues or secreta. BEM is concerned with the measurement and assessment of early, non-adverse, biological alterations in exposed workers to evaluate exposure and/or health risk compared to appropriate references. HS is concerned with periodic medico-physiological examination of exposed workers with the objective of protecting and preventing occupationally related diseases (Zielhuis and Henderson, 1986). EM was shown to be of limited value for assessing the internal dose of a chemical by not taking into account for example toxicokinetic and toxicodynamic processes determining the ultimate fate of xenobiotics in the body. To a certain extent, BM appeared to overcome the problems inherently related to EM. BM assesses the overall exposure to xenobiotics that are present at the workplace through measurement of the appropriate determinant(s) in biological specimens collected from the worker at specific timepoints (ACGIH, 1990). Ideally, not only the relation between exposure and effect is known, but also the toxicokinetic and toxicodynamic interactions linking these two. If these processes are elucidated, quantitative knowledge of a determinant of one of the different monitoring methods allows an assessment either of the level of exposure or of the level of effect (Figure 2.5). For example, the level of urinary mercapturic acid excretion could assess the potential health hazard of an occupational exposure situation (Henderson et al., 1989). In practice, a complete view on the relation between toxicokinetics and toxicodynamics has not been elucidated for a single chemical up to now. Occupational monitoring methods all have their specific values based on their selectivity, sensitivity, validity and logistics and should therefore be used complementary to each other. All methods operate on the continuum from exposure to effect, the limits between which occupational toxicology studies operate.

20 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS

Figure 2.5 Occupational monitoring methods and their relation to exposure versus effect assessment and to toxicokinetic and toxicodynamic processes. (Adapted from Henderson et al., 1989).

Glutathione conjugation products as biomarkers In principle, GSH-conjugation derived metabolites can be used as a biomarker of internal dose. Glutathione (GSH), a tripeptide consisting of the amino acids glycine, cysteine and -glutamine, plays an important role in the detoxification of potentially electrophilic chemicals or metabolites. In contrast, toxification via GSH-conjugation, for example of 1,2dibromoethane, hexachlorobutadiene, benzyl- and allylisothiocyanate has also been reported. β-lyase dependent bioactivation of cysteine-conjugates, derived from the initially formed GSH-conjugates, sometimes resulted in the formation of new reactive intermediates which are responsible for carcinogenic, mutagenic and other toxicological effects (Vermeulen, 1989; Van Welie et al., 1992). The initial step in GSH-conjugation is reaction of the nucleophilic sulphhydryl with electrophilic centers of a chemical. GSH-conjugation is catalysed by a family of glutathione S-transferase (GST) enzymes. A wide range of chemicals can be handled by this enzyme system due to the

N.P.E.VERMEULEN ET AL. 21

Figure 2.6 Schematic representation of the mercapturic acid pathway: GSHconjugation with an electrophilic chemical (RX) and the biosynthesis to a mercapturic acid. E1: glutathione S-transferase, E2: -glutamyltranspeptidase, E3: cysteinylglycinase and aminopeptidase, E4: cysteine conjugate N-acetyltransferase, E5: N-deacetylase.

existence of a large number of isoenzymes with different, though overlapping, substrate selectivity. The final detoxification capacity through GSH and GST enzymes of an organism depends on endogenous factors such as tissue distribution, genetic deficiencies, aging and hormonal influences and on exogenous factors such as sensitivity to inhibition and induction of GSTs (Vermeulen, 1989; Van Welie et al., 1992). GSH-conjugates normally are not excreted unchanged in urine or faeces. Catabolism of the GSH-conjugates results in the formation and excretion of a variety of sulphur containing metabolites, among which thioethers and mercapturic acids (S-substituted N-acetyl-cysteine conjugates) belong to the most important. The mercapturic acid pathway is shown in Figure 2.6. Thioethers in human studies Several years ago, Seutter-Berlage et al. proposed the appearance of thioethers such as mercapturic acids (R-S-R′), mercaptans (R-SH) and disulfides (R-S-S-R′) in urine as an indicator of exposure to potentially alkylating chemicals. The thioether assay is an aselective assay to detect metabolic end-products excreted in urine of (non)occupational exposure to various electrophilic chemicals. It includes three steps, namely: (i) extraction, (ii) alkaline hydrolysis and (iii) derivatization, subsequently followed by spectrophotometric analysis at 412 nm. The thioether assay

22 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS

Figure 2.7 Urinary excretion of thioethers (mmol SH/mol creatinine), of applicators exposed to 3.8 (− −), 9.8 (− −) and 18.9 (− −) mg m−3 8-h TWA (Z+E)-1,3dichloropropene in respiratory air, respectively. Darker shaded areas indicate exposure periods.

was first applied to compare thioether excretions in urine of employees of a chemical plant. Highest thioether excretions were found in rubber workers and radial tyre builders when compared with clerks, plastic monomer mixers and footwear preparers. Recently, urinary thioether excretion was related to the occupational respiratory exposure of applicators in the Dutch flower-bulb culture to 1,3-dichloropropene (DCP) (Van Welie et al., 1991a). Instead of a discrete comparison of thioether excretion with exposed versus non-exposed groups, in this study thioether excretion was related to a continuous scale of airborne DCP concentrations. Significant linear relations between respiratory exposure to DCP and postminus preshift thioether concentration and cumulative thioether excretion were found. The urinary excretion of DCP-thioethers followed first-order elimination kinetics (Figure 2.7) with half-lives of elimination of 8.0±2.5 h (n=5) based on urinary excretion rates and 9.5±3.1 h (n=5) based on creatinine excretion. The elimination half-lives of the thioethers were almost two fold higher when compared to the half-lives of elimination of the mercapturic acids of Z-and E-1,3 dichloropropene. This illustrates the main problem of urinary thioethers, viz. high background levels originating from endogenous or exogenous sources, such as smoking and diet (e.g. horse radish, onion and garlic).

N.P.E.VERMEULEN ET AL. 23

Mercapturic acids in human studies Mercapturic acids, S-substituted N-acetyl-L-cysteine S-conjugates, in urine can be used as biomarkers of internal dose of electrophilic xenobiotics. Mercapturic acids are metabolic end products of GSH-conjugation of various potentially electrophilic chemicals (Figure 2.6). The first mercapturic acids were identified in 1879 as sulphur containing metabolites after administration of bromobenzene to dogs (see references in Vermeulen, 1989). Since then mercapturic acids from many chemicals have been identified and these types of urinary metabolites have been used in biotransformation, biological monitoring and toxicological studies (Vermeulen, 1989; Van Welie et al., 1992). Commercial availability of reference compounds and the development of a number of different analytical techniques attributed to the popularity of mercapturic acids in biological monitoring studies during the last few years. Urinary excretion of the stereoisomeric mercapturic acids of Z- and E-1,3-dichloropropene, a soil fumigant frequently used in agriculture, proved to be a suitable biomarker for the exposure to both isomers in man. Strong correlations were observed between 8-h time weighted average exposure to Z- and E-DCP and complete cumulative excretion of N-acetylS-(Z- and E-3-chloropropenyl-2)-L-cysteine in urine. N-acetyl-S(cyanoethyl)-L-cysteine was proposed as biomarker of exposure to acrylonitrile. The best correlation between uptake of acrylonitrile via the lungs and excretion of the cyanoethyl mercapturic acid in urine was obtained in samples collected between the sixth and the eighth hour after the beginning of exposure (Jakubowoski et al., 1987). The phenyl mercapturic acid of benzene was regarded as a useful biomarker of exposure below 1 ppm of workers in a chemical production plant (Stommel et al., 1989). The use of certain foodstuffs and drugs may also give rise to the excretion of mercapturic acids. Consumption of cabbage and horse radish for example gave rise to increased thioether excretion. Consumption of garlic and onions resulted in the excretion of N-acetyl-S(allyl- and 2-carboxypropyl)-L-cysteine in urine (Van Welie et al., 1992). The hypnotic drug (α-bromo-isovalerylurea also gave rise to the excretion of two diastereomeric α-bromoisovalerylurea mercapturic acid conjugates in urine (Mulders et al., 1993). S-Phenyl mercapturic acid was present in urine of groups of smokers and non-smokers, not exposed to benzene, in concentrations of 4.0±4.0 µg g−1 creatinine (Stommel et al., 1989). Toxicokinetics Knowledge about the toxicokinetics of mercapturic acids is necessary to develop optimal sampling strategies in occupational studies. Urinary excretion rates of mercapturic acids theoretically may reflect the rates of

24 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS

Figure 2.8 Urinary excretion (– –=Z, − −=E) and cumulative excretion (– –=Z, − −=E) of Z- and E-DCP-MA of an applicator due to an 8-h TWA respiratory exposure to 2.32 mg m−3 Z-DCP and 1.73 mg m−3 E-DCP. In (A) the mercapturic acid excretion rate is depicted and in (B) the mercapturic acid excretion based on creatinine excretion.

elimination of the parent compounds from blood and can be used to calculate the (complete) cumulative excretion of mercapturic acids related to exposure. By knowing an individual’s mercapturic acid excretion rate, the contribution to urinary mercapturic acid excretion of the day under study on succeeding day(s) can be calculated. The contributions of previous days of exposure can also be used to correct the mercapturic acid excretion of the exposure day under study. The urinary half-life of elimination is inversely proportional to the elimination rate constant. The urinary halflife of both mercapturic acids of Z- and E-DCP in man was ca. 5 h (Figure 2.8) and they were not significantly different, i.e. 5.0±1.2 h for ZDCP-MA and 4.7±1.3 h for E-DCP-MA. Strong corre-lations (r≥0.93) were observed between respiratory 8 h time weighted average (TWA) exposure to Z- and E-DCP and complete cumulative urinary excretion of Z- and EDCP-MA. There is still a lack of knowledge about the magnitude of the

N.P.E.VERMEULEN ET AL. 25

intra- and inter-individual differences in GSH-conjugation and mercapturic acid excretion. Factors causing these differences are sex, stress, diet, age, enzyme induction and inhibition, pathology and genetic variability. Apart from these factors the presence or absence of glutathione S-transferases (GSTs) or GST activity in different persons is of special interest in relation to urinary mercapturic acid excretion. The most intriguing factor known in this context is the human genetic polymorphism of mu-class GSTs. The GST isoenzyme µ is expressed only in approximately 60% of the human population. Mu-class GST isoenzymes showed a high specific activity towards for example styrene-7,8-oxide and benzo(a)pyrene-4,5dihydrodiol-4,5-oxide and E- and Z-DCP. Genetic polymorphism of muclass GSTs was postulated as a determinant in the excretion of the mercapturic acids of Z- and E-DCP in occupationally exposed applicators. However, between mu-class positive (n=9) and mu-class Table 2.2 Urinary excretion levels, urinary ratios and half-lives of elimination of Zand E-DCP mercapturic acids of mu-class positive and mu-class negative individualsa

a Urinary

excretion level represents the cumulative excretion of Z- and E-DCP-MA in 0–36 h urine, corrected for the time weighted average 8-h exposure to Z- and EDCP. Values are expressed as means±SD for the number of individuals indicated in parentheses. b (mmol mercapturic acid)/(mmol DCP m−3). c Z-DCP-MA/E-DCP-MA d Half-life of elimination

negative (n=3) applicators, neither a difference in urinary half-lives of elimination nor in cumulative excretion of both mercapturic acids of Zand E-DCP was seen (Vos et al., 1991) (Table 2.2). α -Bromoisovalerylurea, a sedative and hypnotic drug, is a racemic drug which is also metabolized by GSH-conjugation. It was proposed as a model substrate to study the pharmacokinetics and stereoselectivity of GSHconjugation in humans. Stereoselective mercapturic acid formation of Rand S-α-bromoisovalerylurea was seen in in vitro studies with purified GST isoenzymes and in vivo in rat and man. In humans, a pronounced stereoselectivity in urinary mercapturic acid excretion was observed. Of an oral dose of R- and S-α-bromoisovalerylurea, 22.5±4.3 and 5.7±1.6% was excreted as mercapturic acid in 24 h, respectively. The half-lives of elimination of both diastereoisomeric mercapturic acids were 1.5±0.4 and 3.1±1.3 h, respectively. Both the pharmacokinetics of α-bromoisovaleryl

26 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS

Figure 2.9 Proposed biotransformation pathway of etridiazol leading to 5-ethoxy-1, 2,4-thiadiazole-3-carboxylic acid (ET-CA) and N-acetyl-S-(ethoxy-1,2,4thiadiazol-3-yl-methyl)-L-cysteine (ET-MA) in rat and humans. Unidentified intermediates are presented between brackets ([...]). GSH: glutathione, MAP: mercapturic acid pathway.

ureas and their stereoselectivity, however, were not found to be different for subjects who were GSH S-transferase class mu deficient and subjects who were not (Mulders et al., 1993). Disposition of etridiazol Etridiazol (Aaterra; 5-ethoxy-3-trichloromethyl-l,2,4-thiadiazole (Figure 2.9)) is an agricultural fungicide used to control phycomycetous fungi in, for example, plants, tomatoes, cucumbers, cauliflowers and celery. Concerning external exposure of applicators (e.g. greenhouse handgunners and foggers) it has been concluded that exposure may occur through inhalation and dermal absorption. For the purpose of the development of a biomonitoring assay disposition studies were performed recently in rats and human volunteers (Van Welie et al., 1991c). Two metabolites, 5-ethoxy-l,2,4-thiadiazole-3-carboxylic acid (ET-CA) and a mercapturic acid, N-acetyl-S-(5-ethoxy-l,2,4-thiadiazol-3-yl-methyl)-Lcysteine (ET-MA) were identified as new metabolites. Based on a preliminary toxicokinetic study, the urinary excretion of the former metabolite amounted to 22±9% of an oral dose of etridiazol (while ET-MA and unchanged etridiazol were less than 1 % of the dose), ET-CA was proposed as a possible biomarker of exposure to this fungicide.

N.P.E.VERMEULEN ET AL. 27

1,1,2-Trichloroethylene The solvent properties of 1,1,2-trichloroethylene (TRI) have resulted in its widespread use in metal degreasing and a wide variety of other industrial applications. TRI has now been in common use for more than 50 years. During this period of time, workers have been exposed to a wide range of concentrations, in some cases for periods of 25 years or longer. This has allowed the compilation of a great data base about the effects of TRI on human health. Moreover, information has been supplemented by numerous studies in experimental animals. Epidemiological studies on more than 15000 individuals with a followup of more than 25 years have shown no evidence of an association between human exposure to TRI and increased incidence of cancer or cancer mortality. However, several of these studies had more or less serious shortcomings. A summary of effects related to TRI and/or TRI-related metabolism is given in Table 2.3. These and other data are taken from Goeptar et al., 1995a. An increased incidence of lung tumors has been reported in female B6C3F1 and male Swiss mice exposed to TRI by inhalation. The effect was not observed in male B6C3F1 nor in female Swiss mice nor in rats. This apparent strain-, sex- and lung-specific response fails to resolve the issue of whether or not TRI is a carcinogenic hazard to man. Mechanistic studies on mouse lung tumor formation have explained the sex and species differences. In this context, chloral formation (Figure 2.10) in Clara cells, containing relatively high cytochrome P-450 concentrations, has been identified to be responsible for the development of mouse lung tumors. Importantly, lung tumors have not been found in humans after long-term occupational exposure in TRI. TRI causes an increase in the incidence of liver cancer in both sexes of B6C3F1 and Swiss mice following either gavage or inhalatory exposure, but not in NMRI and Ha: ICR mice nor in rats. A rodent specific link between peroxisome proliferation, DNA synthesis, inhibition of intercellular communication and cancer (Table 2.3) suggests that these responses are the basis of the hepatocarcinogenicity induced by TRI. The identification of TCA in cancer bioassays as the responsible metabolite for these effects confirmed this hypothesis. However, when TCA was administered to both rats and mice, liver cancer was only observed in mice and not in rats. The reason for this species selectivity in liver effects is explained by the kinetic behavior of TRI and TCA in rodents. Both rats and mice have a considerable capacity to metabolize TRI to TCA and TCE, the maximal capacities being closely related to the relative surface areas rather than to their body weights. Oxidative metabolism of TRI in rats is linearly related to dose at lower dose levels, but it becomes saturated at higher dose levels. Thus, an important difference between rats and mice is the lower saturation

References: see review Goeptar et al., 1995b. n.d.: not determined.

a

Table 2.3 Reported toxic effects related to TRI and/or TRI-derived metabolites

28 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS

N.P.E.VERMEULEN ET AL. 29

Figure 2.10 Oxidative metabolism of TRI in the rodent and mammalian liver and the formation of metabolites which are excreted in the urine.

concentration in the former species. The relevance of the mechanisms of liver tumor formation in B6C3F1 and Swiss mice for humans exposed to TRI has been assessed in studies comparing metabolic rates in mice, rats and humans. In contrast to the rat, the oxidative metabolism of TRI to TCA in humans is not limited by saturation. In this respect, humans resemble the mouse and might be able to produce sufficient TCA to induce peroxisome proliferation and consequently liver cancer. However, there are significant differences between mice and humans. First, humans metabolize approximately 60 times less TRI on a body weight basis than mice at similar exposure levels. Second, TCA has been shown to induce peroxisome proliferation in mouse hepatocytes but not in human hepatocytes (Table 2.3). Consequently, the combination of extensive oxidative metabolism of TRI to TCA and the ability of TCA to induce peroxisome proliferation appear to be unique to B6C3F1 and Swiss mice. TRI-induced renal toxicity and tumors were found in Sprague-Dawley, Fischer 344 and Osborne-Mendel rats. These nephrocarcinogenic effects of TRI were specific to male rats and were not seen in female rats nor in mice of either sex. 1,2-DCV-Cys, formed from TRI via the mercapturic acid pathway, has been identified as a likely metabolite involved in the observed renal toxicity and probably also in renal carcinogenicity in rats. TRI is metabolized by a minor pathway involving initial hepatic GSH-conjugation of TRI. The resulting DCV-G is further metabolized (Figure 2.11) and excreted in urine as two regioisomeric mercapturic acids, namely vicinal 1, 2-DCV-Nac and geminal 2,2-DCV-Nac (Figure 2.11). 1,2-DCV-Cys (the precursors of 1,2-DCV-Nac) is a substrate for the renal L-cysteine S-

30 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS

Figure 2.11 Possible routes of metabolism of S-(1,2-dichlorovinyl)glutathione (1,2DCV-G). Steps are catalyzed by (a) -glutamyltransferase; (b) cysteinylglycine dipeptidase; (c) L-cysteine S-conjugate β-lyase; (d) L-cysteine S-conjugate Nacetyltransferase; (e) acylase.

conjugate β-lyase and it is more mutagenic and cytotoxic than 2,2-DCVCys (the precursors of 2,2-DCV-Nac). The bioactivation of 1,2-DCV-Cys is without a doubt a crucial step in the onset of nephrotoxicity in the rat, although the precise biological mechanisms by which these metabolites exert their nephrocarcinogenic effects are not yet fully understood. A key aspect in the onset of nephrocarcinogenicity in rats, however, is that it will not occur in the absence of nephrotoxicity. This suggests that the alkylating effects of the reactive metabolites (most likely thioketenes) derived from bioactivation of 1,2-DCV-Cys by β-lyase may not be sufficient to cause kidney tumors. The specific activity of β-lyase, the key enzyme involved in the bioactivation of DCV-Cys isomers, is similar in humans to that in the mouse and only 10% of that in the rat. Moreover, human TRI metabolism via the mercapturic acid pathway resembles that of the mouse. It is, therefore, questionable whether humans are able to produce sufficient DCV-Cys isomers from TRI to cause first nephrotoxicity and then nephrocarcinogenicity. An important finding is also that the occurrence of nephrotoxicity and

N.P.E.VERMEULEN ET AL. 31

nephrocarcinogenicity in the male rat is dose-dependent. More specifically, cytotoxic kidney damage is a feature of high continuous exposure to TRI over prolonged periods of time. This is unlikely to occur in humans during occupational exposure. In fact, TRI has been found not to be nephrotoxic in humans chronically exposed to low levels of TRI (50 mg m−3). Consequently, it is unlikely that the renal tumors which are seen in rats at nephrotoxic dose levels of TRI and which are related to β-lyase mediated bioactivation of 1,2-DCV-Cys, are relevant to human health hazards at reasonably foreseeable levels of exposure.

Physiologically based toxicokinetic modeling of 1,3butadiene Physiologically based pharmaco(toxico)-kinetic models differ from the conventional compartmental models in that they are based to a large extent on the actual physiology of the organism. Instead of compartments defined largely by the experimental data themselves, actual organ and tissue groups are used with weights and blood flows from the literature (Bischof and Brown, 1966). Instead of composite rate constants determined by fitting the actual experimental data, physical-chemical and biochemical constants of the compound are used. The result is a mode which predicts the qualitative behavior of the experimental time course without being based on it. Refinements of the model to incorporate additional insights gained from comparison with experimental data yields a model which can be used for quantitative extrapolations well beyond the range of experiments. In recent years several PBTK- and PBPK-models have been published: for methylene chloride, see Andersen et al., 1987; for a review see Leung et al., 1988; for 1,3-butadiene, see Evelo et al., 1993. The development of a PBTK/PBPK model can be divided into a number of steps: (a) inventory of physiological and toxicological behaviour of the compound, (b) mathematical description of the biochemical/(patho) physiological processes involved, (c) parameterization of the mathematical descriptions, (d) the construction of the model, (e) refinement and validation of the model and (f) use of the predictions and risk assessment. As an illustrative example of this approach the recently described PBTKmodeling of 1,3-butadiene disposition and toxicity might be used (Evelo et al., 1993). 1,3-Butadiene used for the production of styrene-butadiene rubber, is known amongst others to cause lung carcinogenicity. In the rat the carcinogenicity of 1,3-butadiene is less pronounced while the evidence for human carcinogenicity is inconclusive, Monoand di-epoxy-butadiene are reactive metabolites held responsible for this effect. Butadiene monoxide is formed by microsomal fractions of the lung and liver of several

32 TOXICOKINETICS AND BIODISPOSITION OF INDUSTRIAL CHEMICALS

Figure 2.12 Physiologically based toxicokinetic model for description of butadiene distribution and metabolism in mice, rats and humans. Gas exchange occurs in the alveoli of the lung. Metabolism occurs in both the alveolar and bronchial areas of the lung and in the liver. Metabolic activity in the three other compartments is ignored (Evelo et al., 1993).

species. There are, however, large interspecies differences in the lung vs liver activities: mice>rats>humans/monkeys. The PBTK model used to describe butadiene distribution and metabolism in mice, rats and humans is shown in Figure 2.12. Gas exchange is supposed to occur in the alveoli of the lung and metabolism in both the alveolar and bronchial areas of the lung and in the liver. By using the experimentally determined or estimated species-selective parameters for

N.P.E.VERMEULEN ET AL. 33

volumes, masses and blood flows of different organs, partition coefficients of 1,3-butadiene between blood and organs/tissues and for metabolic capacities in liver and lung (bronchial and alveolar areas), accurate dosedependent simulations were performed for the uptake of 1,3-butadiene in mice and rats in gas-closed chambers. Moreover, with the resulting model the relative importance of lung metabolism as compared to metabolism in the liver was predicted for the three different species. Lung metabolism appeared to be much more important than liver metabolism in mice, this in contrast to the situation in the rat and humans. Moreover, at low exposure concentrations the relative importance of lung metabolism was predicted to increase in mice as a result of diminished saturation of metabolism in this species. It was concluded that the observed species differences in lung vs liver metabolism of 1,3-butadiene (mice>rat>human) and the tendency towards increased lung metabolism at low doses might rationalize the observed species differences in the lung carcinogenicity of 1,3-butadiene and this knowledge should be useful in the in vivo extrapolation from high dose to low dose risk assessments within one species as well as in interspecies risk assessment extrapolations. Conclusions In conclusion, a profound knowledge of the biodisposition and the toxicokinetics of a toxic or potentially toxic chemical is of utmost importance to the design and interpretation of laboratory assessments of toxicity, to explain interspecies differences in toxicities and to extrapolate more reliably from animal experiments to man in the process of risk assessment. This also holds true for the design for proper biological monitoring procedures and for the interpretation of the results in terms of potential health risks of exposure to chemicals. Apart from traditional compartment-based toxicokinetic approaches, more recent physiologicallybased toxicokinetics modeling approaches have distinct advantages for the above-mentioned purposes. References ACGIH, 1990, in 1990–1991 Threshold limit values for chemical substances and physical agents and biological exposure indices, American Conference of Governmental Industrial Hygienists, No. 0205. ANDERSEN, M.E., CLEWELL, H.J., GARGAS, M.L., SMITH, F.A. and REITZ, R.H., 1987, Physiologically-based pharmacokinetics and the risk assessment for methylene chloride, Toxicol. Appl. Pharmacol., 87, 185–205. ARIENS, E.J. and SIMONIS, M.A., 1980, in BREIMER, D.D. (Ed.) Towards better Safety of Drugs and Pharmaceutical Products, Amsterdam: Elsevier Biomedical Press.

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BISCHOF, K.B. and BROWN, R.G., 1966, Drug distribution in mammals, Chem. Eng. Prog. Symp. Ser., 62(66), 33–45. CCRX 1989, in Metingen van radioactiviteit en xenobiotische stoffen in het biologische milieu in Nederland 1989 (in Dutch with English summary), Coördinatie-commissie voor de metingen van radioactiviteit en xenobiotische stoffen, Bilthoven: RIVM. COMMANDEUR, J.N.M. and VERMEULEN, N.P.E. 1991, Molecular and biochemical mechanism of chemically induced nephrotoxicity: a review, Chem. Res. Toxicol., 3, 171–94. EVELO, C.T.A., OOSTENDORP, J.G.M., TEN BERGE, W.F. and BORM, P.J. A., 1993, Physiologically based toxicokinetic modeling of 1,3-butadiene lung metabolism in mice becomes more important at low doses. Environ. Hlth Perspect., 101(6), 496–502 (no. 24). FARMER, P.B., NEUMANN, H.-G. and HENSCHLER, D., 1987, Estimation of exposure of man to substances reacting covalently with macromolecules, Arch. Toxicol, 60, 251–60. GOEPTAR, A.R., COMMANDEUR, J.N.M., OMMEN, B.VAN, BLADEREN, P.J. VAN and VERMEULEN, N.P.E. 1995a, The metabolism and kinetics of trichloroethylene in relation to toxicity and carcinogenicity. Relevance of the Mercapturic Acid Pathway, Chem. Res. Toxicol, 8, 3–21. GOEPTAR, A.R., SCHEERENS, H. and VERMEULEN, N.P.E., 1995b, Oxygen and xenobiotic reductase activities of cytochrome P450, Crit. Rev. Toxicol., 25, 25–65. GUENGERICH, F.P., 1994, Catalytic selectivity of human cytochrome P450 enzymes: relevance to drug metabolism and toxicity, Toxicol. Lett., 70, 133–8. HENDERSON, R.F., BECHTOLD, W.E., BOND, J.A. and SUN, J.D., 1989, The use of biological markers in toxicology, Crit. Rev. Toxicol, 20, 65–82. HOOK, J.B. and HEWITT, W.R., 1986, Toxic responses of the kidney, in Klaassen, C.D., Doull, J. and Amdur, M.O. (Eds) Casarett and Doull’s Toxicology, pp. 310–29, New York: Macmillan. JAKUBOWOSKI, M., LINHART, I., PIELAS, G. and KOPECKY, J., 1987, 2Cyanoethylmercapturic acid (CEMA) in the urine as a possible indicator of exposure to acrylonitrile, Brit. J. Ind. Med., 44, 834–40. KOYMANS, L., DONNÉ-OP DEN KELDER, G.M., TE KOPPELE, J.M. and VERMEULEN, N.P.E., 1993, Cytochromes P450: their active-site structure and mechanism of oxidation, Drug Metab. Rev., 25, 325–87. LEUNG, H.W., Ku, R.H., PAUSTENBACH, D.J. and ANDERSEN, M.E., 1988, A physiologically-based pharmacokinetic model for 2,3,7,8-tetrachlorodibenzo-pdioxin in C57BL/6J and DBA/2J mice, Toxicol. Lett., 42, 15–28. MULDERS, T.M.T., VENIZELOS, V., SCHOEMAKER, R., COHEN, A.F., BREIMER, D.D. and MULDER, G.J., 1993, Characterization of glutathione conjugation in humans: stereoselectivity in plasma elimination pharmacokinetics and urinary excretion of (R)- and (S)-2-bromoisovalerylurea in healthy volunteers. Clin. Phar. Ther., 53, 49–58. POTT, P. 1795, Chirurgical observations relative to the cataract, the polypus of the nose, the cancer of the scrotum, the different kinds of ruptures and the mortification of the toes and feet, in Haes, Clarke and Collins (Eds) National Cancer Institute Monograph, 1962, Vol 10, pp. 7–13, London.

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SEUTTER-BERLAGE, F., VAN DORP, H.L., KOSSE, H.G.J. and HENDERSON, P.T.H., 1979, Urinary mercapturic acid excretion as a biological parameter of exposure to alkylating agents, Int. Arch. Occup. Environ. Hlth, 39, 45–51. SIPES, I.G. and GANDOLFI, A.J., 1986, Biotransformation of chemicals, in Klaassen, C.D., Doull, J. and Amour, M.O. (Eds) Casarett and Doull’s Toxicology, pp. 64–98, New York: Macmillan. STOMMEL, P., MÜLLER, G., STÜCKER, W., VERKOYEN, C., SCHÖBEL, S. and NORPOTH, K., 1989, Determination of S-phenylmercapturic acid in the urine an improvement in the biological monitoring of benzene exposure, Carcinogenesis, 10, 279–82. VAN WELIE, R.T.H., VAN MARREWIJK C.M., DE WOLFF, F.A. and VERMEULEN, N.P.E., 1991a, Thioether excretion in urine of applicators exposed to 1,3-dichloropropene: a comparison with urinary mercapturic acid excretion, Brit. J. Ind. Med., 48, 492–8. VAN WELIE, R.T.H., VAN DUYN, P., BROUWER, E.J., VAN HEMMEN, J.J. and VERMEULEN, N.P.E., 1991b, Inhalation exposure to 1.3dichloropropene in the Dutch flower-bulb culture. Part II. Biological monitoring by measurement of urinary excretion of two mercapturic acid metabolites, Arch. Environ. Contam. Toxicol, 20, 6–12. VAN WELIE R.T.H., MENSERT, R.,, VAN DUYN, P. and VERMEULEN, N.P. E. 1991c, Identification and quantitative determination of a carboxylic and a mercapturic acid metabolite of etridiazole in urine of rat and man. Potential tools for biological monitoring. Arch. Toxicol., 65, 625–32. VAN WELIE, R.T.H., VAN DIJCK, R.G.J.M., VERMEULEN, N.P.E. and VAN SITTERT, N.J., 1992, Mercapturic acids, protein adducts, and DNA adducts as biomarkers of electrophilic chemicals, Crit. Rev. Toxicol., 22, 271–306. VERMEULEN, N.P.E., 1989, Analysis of mercapturic acids as a tool in biotransformation, biomonitoring and toxicological studies. TiPS, 10, 177–81. VERMEULEN, N.P.E., VAN DER STRAAT, R., TE KOPPELE, J.M., BALDEW, G.S., COMMANDEUR. J.N.M., HAENEN, G.R.M.M., KOYMANS, L. and VAN WELIE, R.T.H., 1990, Molecular mechanisms in toxicology and drug design, in Claassen, V. (Ed.) Vol. 13, pp. 253–71, Trends in Drug Research, Amsterdam: Elsevier. Vos, R.M.E., VAN WELIE, R.T.H., PETERS, W.H.M., EVELO, C.T.A., BOOGAARDS, J.J.P., VERMEULEN, N.P.E. and VAN BLADEREN, P.J., 1991, Genetic deficiency of human class mu glutathione S-transferase isoenzymes in relation to the urinary excretion of the mercapturic adds of Zand E-1,3-dichloropropene. Arch. Toxicol., 65, 95–9. WARHOLM, M., JENSSON, H., TAHIR, M.K. and MANNERVIK, B., 1986, Purification and characterization of three distinct glutathione S-transferases from mouse liver, Biochemistry., 25, 4119–25. ZIELHUIS, R.L. and HENDERSON, P.TH., 1986, Definitions of monitoring activities and their relevance for the practice of occupational health, Int. Arch. Occup. Environ. Hlth, 57, 249–57.

3 Metabolic Activation of Industrial Chemicals and Implications for Toxicity GERARD J.MULDER Leiden University, Leiden

Introduction In the toxicity of industrial chemicals bioactivation (Anders, 1985) plays an important role. Obviously, its importance depends on the structure of the chemical as well as the toxic effect considered. Thus, inorganic compounds in general will not require bioactivation: metal salts or oxides will usually cause toxicity in the form in which they are taken up. However, even these chemicals may require further metabolism for maximum toxicity in the body: inorganic mercury may be converted to an organic form (methylmercury), and nitrate may be reduced to nitrite. It is also possible that in vivo complexes are being formed, such as between heavy metals ions and the protein, metallothionein, which may be more toxic (or cause more organ-selective toxicity) than the original, uncomplexed compound (Wang et al., 1993). Bioactivation thus mostly concerns the conversion of organic chemicals to more toxic products. On one hand this may result in stable metabolites that better fit a receptor binding site, resulting in (in principle) reversible interactions (Mulder, 1992). On the other hand, the metabolites may be quite reactive, resulting in essentially irreversible effects which are of particular concern when they can escape correction, such as neoplasms or sensitization. Mechanisms of bioactivation Industrial chemicals have widely different structures. Often the preparations used contain a variable degree of impurities, or are mixtures. In this chapter only the toxicity of pure chemicals will be discussed; obviously when several compounds are present at the same time in a reaction mix or a commercial product, the final toxicity may be the result of complex interactions between the substituents, which may cause the toxicity to be more severe (but also much less serious) than expected.

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The bioactivation to reactive intermediates by oxidative, cytochrome P450-mediated metabolism has been extensively studied. So much so, that it is often overlooked that conjugation reactions may similarly convert stable compounds into reactive, electrophilic metabolites (Anders and Dekant, 1994). This is of some practical importance, because many rapid in vitro toxicity screening tests, e.g. for genotoxicity, include only oxidative biotransformation capacity (microsomal fractions plus NADPH). In such screening systems the possibility that, for example, glucuronidation, sulfation or glutathione conjugation may activate a chemical is not assessed. Examples of bioactivation of industrial chemicals by glutathione conjugation are various halogenated hydrocarbons, while in 2naphthylamine toxicity glucuronidation may play a role. All in all, however, little information is available on the role of conjugation. As a consequence, it is unclear at present whether conjugation reactions are of major concern for bioactivation of industrial chemicals in general. It certainly seems worth while for reasons more than just scientific curiosity to include conjugation reactions in test systems. This can be done by using, for example, intact hepatocytes (or other cells), or by using a mix of cosubstrates for conjugation in combination with an S9 fraction (consisting of both cytosol and microsomal fraction). UDP glucuronic acid, a sulfate activating system, glutathione, acetyl-CoA and S-adenosylmethionine would cover the major conjugation reactions. A role of bioactivation in the toxicity of many chemicals has been demonstrated. Chemical groups that often are involved in mutagenic or carcinogenic effects have been identified (‘alerting groups’). However, as yet it is still impossible to predict with certainty the carcinogencity of a compound based only on its chemical structure, although a panel of experts can make quite good guesses (Wachsman et al., 1993). In this chapter some of the major issues will be illustrated by the examples vinyl chloride, styrene (versus styrene oxide), benzene, dichloromethane, chloroform, 1,2-dibromoethane and 2-naphthylamine. Vinyl chloride High exposure of workers to vinyl chloride in the past has led to the realization that it may cause neoplasms in man, in particular haemangiosarcomas in the liver. Vinyl chloride is a genotoxic compound that acts as initiator of various types of tumors (Swaen et al., 1987). The major routes of bioactivation of vinyl chloride are shown in Figure 3.1. The most important first step is oxidation by (a) cytochrome P450 species, resulting in a rather reactive epoxide, which readily rearranges to chloroacetaldehyde. This may bind to DNA bases, especially the N6 of adenosine or the N4 of cytidine, yielding N-ethenoadducts. Glutathione provides protectionbecause it traps the reactive intermediates

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Figure 3.1 Bioactivation of vinylchloride.

formed from vinyl chloride. Furthermetabolism of such conjugates leads to urinary products that can be used tomonitor vinyl chloride exposure in workers (Guengerich, 1992). The compound is mutagenic in many in vitro test systems, which require bioactivation by a microsomal preparation with co-factors for cytochrome P450. Whether other toxic effects that have been associated with vinyl chloride exposure in man, such as Raynauds syndrome or acro-osteolysis, also require bioactivation of vinyl chloride is unknown. In addition to its DNA adduct forming capacity, vinyl chloride also binds covalently to thiol groups in proteins. It is conceivable that such binding in specific cell types might lead to non-carcinogenic defects in organ functions. Styrene and styrene oxide Styrene metabolism and bioactivation are very similar to that of vinyl chloride: epoxidation by cytochrome P450 is the pathway of toxification (Figure 3.2). It can be detoxified by epoxide hydrolase and glutathione transferase activity. Mandelic acid excretion in urine can be used for exposure monitoring in man. Styrene oxide is a direct mutagen in several in vitro mutagenesis systems and it readily reacts with DNA in vitro. However, when animals are exposed to styrene in vivo very little if any DNA binding is observed. Moreover, styrene is not carcinogenic in animal experiments, although it is a (weak) mutagen in vitro, after bioactivation (Bond, 1989; Ecetoc, 1992). The explanation most likely is that the styrene

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Figure 3.2 Bioactivation of styrene.

Figure 3.3 Bioactivation of chloroform.

oxide, generated in vivo inside a cell is such a good substrate for the phase 2 enzymes, epoxide hydrolase and glutathione transferase, that virtually immediately upon its synthesis, it is further metabolized. Thus, presumably the build-up of an effective concentration in vivo is prevented. Whether other toxicity of styrene in, for example, oesophagus, stomach or forestomach is related to covalent binding of styrene oxide to protein thiol groups in those tissues is unclear at present. Styrene is an example of a compound of which the metabolism completely goes through a reactive intermediate (the epoxide); yet it does not cause the cancer that might be expected from its highly mutagenic metabolite. Accumulation of enough of this epoxide inside the cells for a detectable genotoxic effect may require a dose which is acutely toxic, and therefore can never be tested. Chloroform Chloroform is acutely toxic in the liver and the kidney. This is the result of formation of a reactive intermediate (Figure 3.3), phosgene, which binds

40 METABOLIC ACTIVATION OF INDUSTRIAL CHEMICALS

avidly to thiol and amine groups in protein. In mice the kidney toxicity is much more pronounced in males than in females; this sex-difference is due to the much higher activity of the bioactivating cytochrome P450 species in male mouse kidney than in the females (Pohl et al., 1984). Chloroform also increased the tumor incidence in the liver and kidney in some experiments (Reitz et al., 1990), at dose levels which damaged these organs. However, there are no indications of mutagenicity or genotoxicity in in vitro or animal in vivo systems. Therefore, most likely the increased tumor frequency in animals is due to tissue toxicity, leading to increased cell turnover and a mitogenic stimulus. This is an important distinction, at least in some countries such as The Netherlands, because for such chemicals a threshold approach is allowed, whereas for initiating chemicals a linear extrapolation for carcinogenic risk is used. Benzene Benzene presents something of a mystery in the evaluation of its toxicity mechanism (Swaen et al., 1989). Exposure to high levels of benzene has been associated with leukaemia in man. However, in vitro it shows little genotoxicity, and it hardly generates DNA adducts when it is given even at high dose to animals. A candidate for DNA damage could have been the 1, 4-dihy-droxybenzene (hydroquinone) metabolite, which, however, does not form DNA adducts readily. Recently a ring-opened metabolite, the trans,trans-muconic dialdehyde has been proposed as a possible reactive metabolite of benzene (Figure 3.4). Whether it really plays a role in benzene toxicity is unclear as yet (Kline et al., 1993). Dichloromethane Dichloromethane can be metabolized by two pathways, an oxidative and a conjugative route. Oxidation catalyzed by P450 yields carbon monoxide (Figure 3.5). The glutathione pathway generates a reactive intermediate, which is mutagenic and has been implicated in the hepatocarcinogenic effect of dichloromethane in mice. It could be shown that the human liver has a negligible activity of the glutathione transferase involved, so that the risk for hepatocarcinogenesis in man is virtually non-existent (Green et al., 1988; Reitz et al., 1989; Dankovic and Bailer, 1994). This example illustrates how insight into the mechanism of bioactivation enables a more reliable species extrapolation in terms of hazard and risk. 1,2-Dibromoethane This compound can be conjugated with glutathione to form a reactive thiiranium ion which forms adducts with DNA. This is the reason for the

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Figure 3.4 Possible route of bioactivation of benzene.

Figure 3.5 Bioactivation of dichloromethane.

carcinogenic and mutagenic effects of 1,2-dibromoethane (Inskeep et al., 1986). 2-Naphthylamine 2-Naphthylamine causes bladder tumors in the dog and man, but not in mice and rats. The most likely cause is a complicated interplay between glucuroni dation and urinary pH. In all four species 2-naphthylamine is Nhydroxylated and subsequently N-glucuronidated. The resulting metabolite

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is excreted in urine. In man and dog the urine is slightly acidic, while in rat and mouse it is slightly alkaline. Under acidic conditions the glucuronide is hydrolyzed to generate the hydroxylamine in the bladder. In this case glucuronidation is not a bioactivation, but rather a targeting biotransformation: in man and dog the carcinogenic metabolite is targeted to the bladder, due to the (necessary!) acidic local pH (Kadlubar et al., 1981). Conclusions The above illustrates the importance of bioactivation in toxicity of industrial chemicals. Is it possible to predict bioactivation from the structure? As outlined above, in some cases the compound contains structural elements which make bioactivation to a reactive intermediate quite likely. Whether it does play a role in toxicity then is still uncertain. Test systems to detect reactive intermediates depend on, for example, the availability of the radiolabeled compound; in fact, a very high specific radioactivity is required to detect low levels of binding. Alternatively, radiolabelled glutathione can be used for those reactive intermediates that readily bind to the thiol group of glutathione (Mulder and Le, 1988). Whether such systems can pick up every relevant toxic reactive intermediate remains to be seen. For extrapolation of one species to the other it is important to have insight into the metabolite that is responsible for the toxicity. Therefore, it is more than just of academic interest to know the mechanism of toxicity in safety assessment of industrial chemicals. Unfortunately, it is often not easy to establish such a mechanism beyond reasonable doubt: it may require too many rats to feel comfortable about it if we would have to do this for every chemical used industrially! References ANDERS, M.W. (Ed.), 1985, Bioactivation of Foreign Compounds, Orlando, FL: Academic Press. ANDERS, M.W. and DEKANT, W., 1994, Conjugation-dependent Carcinogenicity and Toxicity of Foreign Compounds, Orlando, FL: Academic Press. BOND, J.A., 1989, Review of the toxicology of styrene, CRC Crit. Rev. Toxicol 19, 227–49. DANKOVIC, D.A. and BAILER, A.J., 1994, The impact of exercise and intersubject variability on dose estimates for dichloromethane derived from a physiologically based pharmacokinetic model, Fund. Appl. Toxicol, 22, 20–5. ECETOC, 1992, Technical report No. 52, Styrene toxicology. Investigations on the potential for carcinogenicity, Brussels: Ecetoc.

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GREEN, T., PROVAN, W.M., COLLINGE, D.C. and GUEST, A.E., 1988, Macro molecular interactions of inhaled methylene chloride in rats and mice, Toxicol. Appl. Pharmacol, 93, 1–10. GUENGERICH, F.R., 1992, Roles of the vinylchloride oxidation products 1chlorooxirane and 2-chloroacetaldehyde in the in vitro formation of etheno adducts of nucleic acid bases, Chem. Res. Toxicol, 5, 2–5. INSKEEP, P.B., KOGA, N.K., CMARIK, J.L. and GUENGERICH, F.P., 1986, Covalent binding of 1,2-dihaloalkanes to DNA, Cancer Res., 46, 2839–44. KADLUBAR, F.F., UNRUH, L.E., FLAMMANG, T.J., SPARKS, D., MITCHUM, R.K. and MULDER, G.J., 1981, Alteration of urinary levels of the carcinogen, N-hydroxy-2-naphthylamine, and its N-glucuronide in the rat by control of urinary pH, inhibition of metabolic sulfation, and changes in biliary excretion, Chem.-Biol. Interact. 33, 129–47. KLINE, S.A., ROBERTSON, J.F., GROTZ, V.L., GOLDSTEIN, B.D. and WITZ, G., 1993, Identification of 6-hydroxy-trans,trans-2,4-hexadienoic acid, a novel ring-opened urinary metabolite of benzene, Environm. Hlth Perspect., 101, 310–12. MULDER, G.J., 1992, Pharmacological effects of drug conjugates: is morphine 6glucuronide an exception? Trends Pharmacol. Sci., 13, 302–4. MULDER, G.J. and LE, C.T., 1988, A rapid simple in vitro screening test to detect reactive intermediates of xenobiotics. Toxicol. In Vitro, 2, 225–30. POHL, L.R., GEORGE, J.W. and SATOH, H., 1984, Strain and sex differences in chloroform-induced nephrotoxicity. Drug Metab. Disposit., 12, 304–7. REITZ, R.H., MENDRALA, A.L. and GUENGERICH, F.P., 1989, In vitro metabo-lism of methylene chloride in human and animal tissues, Toxicol. Appl. Pharmacol, 97, 230–46. REITZ, R.H., MENDRALA, A.L. and CONOLLY, R.B., 1990, Estimating the risk of liver cancer associated with human exposures to chloroform using PbPK modeling, Toxicol. Appl. Pharmacol., 105, 443–59. SWAEN, G.M.H. et al., 1987, A scientific basis for the risk assessment of vinyl chloride, Regul. Toxicol. Pharmacol, 7, 120–7. SWAEN, G.M.H. et al., 1989, Carcinogenic risk assessment of benzene in outdoor air, Regul. Toxicol. Pharmacol., 9, 175–85. WACHSMAN, J.T., BRISTOL, D.W., SPALDING, J., SHELBY, M. and TENNANT, R.W., 1993, Predicting chemical carcinogenesis in rodents, Environm. Hlth Perspect., 101, 444–5. WANG, X.P., CHAN, H.M., GOYER, R.A. and CHERIAN, M.G., 1993, Nephrotoxicity of repeated injections of cadmium-metallothionein in rats, Toxicol. Appl. Pharmacol., 119, 11–16.

4 Sizing up the Problem of Exposure Extrapolation: New Directions in Allometric Scaling D.BRUCE CAMPBELL Director International Scientific Affairs, Servier Research and Development, Slough

Introduction The evaluation of the safety of industrial chemicals requires the administration of a range of doses to test animals over periods of time and the extrapolation in some meaningful way to man. Various risk assessment models have been suggested which attempt to measure an uncertainty or safety factor which can be used to extrapolate to man to obtain an acceptable daily intake (ADI) (Dourson and Stara, 1983). Other approaches are also used, such as benchmark dose, the smallest dose which produces a statistical increase in toxicity over the background level (Crump, 1984), or more frequently the LOEL, the lowest observed dose which produces an adverse effect, and NOEL, the highest dose at which no adverse effect is observed. There are difficulties in the interpretation of these exposure margins since there is often little information on: (1) the slope or intensity of the effect, (2) species differences in the sensitivity, (3) the possibility of cumulative or irreversible toxicities, etc. But perhaps the most important weakness in these estimates is the lack of knowledge of the actual circulating levels of the chemical(s) in the different species. This problem is particularly pertinent for industrial chemicals and environmental pollutants where it may be unethical to administer doses of these compounds to volunteers which are sufficiently high to measure the kinetics. It is of special concern since it is well known that there are large interspecies differences in the clearance of chemicals and that comparison of doses in animals, expressed simply in terms of mg kg−1, provides little information as to the actual exposure likely to occur. This is not surprising since small animals have relatively faster blood flow and larger organs than man when expressed as a percentage of body weight, and consequently clearance is more rapid and circulating levels of the administered compound are lower than could be expected during toxicity testing (Campbell and Ings, 1988). However since most mammals share similar physiological and biochemical actions these differences in physiological rates and sizes for

D.BRUCE CAMPBELL 45

most processes in the mammalian body have been shown to be proportional to the body weight of the animal (Adolph, 1949; Calabrese, 1983; Peters, 1983; Chappell and Mordenti, 1991) and can be related by allometry, a word from the Greek meaning the measurement (metry) of changing size (allo). It has been shown that blood flow, organ size, metabolic and respiratory rate, and many other physiological and anatomical variables are related by the general allometric equation (Boxenbaum, 1982b): (4.1) where Y is the function to be measured, W the body weight of the animal, a the coefficient and b the exponent. For mammals, whilst a is different for each function, b is approximately 0.6–0.8 for rates, flows and clearances, 1. 0 for volumes and organ sizes, and 0.25 for cycles and times. Thus metabolic rate can be calculated from 7.0·W0.75, liver blood flow from 37·W0.85, blood weight from 0.055·W0.99, and respiratory rate from 0. 019·W0.26. Since the blood flows and the weights of the liver and kidney, the two major organs of elimination, can be similarly allometrically scaled, it follows that the same formula could in principle be used for extrapolation of the clearance of chemicals between species. In the past there has been much discussion on the possibility of predicting human kinetics and distribution from animal data, using allometry. For industrial chemicals relatively complex physiological models have been constructed using this knowledge of relative blood flows and organ size to predict what levels of exposure could be expected in man (Andersen et al., 1984), but little work has been published on comparative interspecies clearances which will dictate the circulating levels. For drugs, on the other hand, a number of reports have been published on the rationale for the use of allometric scaling of kinetics (Dedrick, 1973; Boxenbaum, 1982b, 1984, 1986; Mordenti, 1985, 1986; Sawada et al., 1985; Chappell and Mordenti, 1991) but many have been concerned with its theoretical aspects rather than with its practical use for prediction. When scaling has been used, the predictions have not always been accurate, and the method has therefore not had wide usage. This is unfortunate since the ability to predict what will be the blood levels in man, without the need to administer the compound, can potentially have many advantages in drug development and in the safety testing of industrial chemicals where dosing volunteers is often unacceptable. Methods A meta-analysis of the papers related to this subject has been made from those published over the last 20 years. Data before this have largely been rejected due to the poor design of the studies or lack of analytical

46 SIZING UP THE PROBLEM OF EXPOSURE EXTRAPOLATION

precision. In the main the data have come from drugs but the same general considerations would hold for environmental chemicals. Wherever possible the only compounds included in the analysis have been those where unbound clearance after systemic administration has been reported, unless it has been shown that there are little interspecies differences in protein binding or that absorption is known to be complete in all the animals. In the past these provisos have not always been met, leading to incorrect interpretation of the data. In most reports the allometric scaling has used results from at least four species but in some cases up to 11 have been included. Practically this would involve an enormous resource and would be difficult when many compounds are being investigated. For this analysis it has been assumed that only one species will initially be used and the aim of this analysis was to find which single species would provide the best prediction of clearance compared to that found in man. Three methods have been used using data, wherever possible, from mouse, rat, rabbit, dog and monkey (macaques) in a total of 60 compounds, with human unbound clearances ranging from 4 to 150 909 ml min−1. Simple allometric equation Figure 4.1 shows a typical allometric relationship for the clearance of the anticancer drug, fotemustine, showing that equation (4.1) can be made linear for the determination of the variables by logarithmically transforming the body weight (W) and clearances (CL), as shown in equation (4.2) where the exponent b can be calculated from the slope of the linear regression. (4.2) From this analysis of all the available papers, where this has been undertaken with more than four species using data taken from 29 compounds, it was possible to show that the mean exponent (b) is approximately 0.70±0.15 for unbound clearance, but with a range of 0.92– 0.28. This mean value is to be expected since it is comparable to the exponent for the allometric equation relating physiological rates and clearances to weight as for metabolic rate, body surface area, hepatic and renal blood flow, etc. Therefore it would seem that even without a specific knowledge of the clearance in a number of different species, it could be assumed that the exponent of 0.7 is a common factor for all chemicals, if it has not been previously determined. The coefficient a can subsequently be determined for each compound from only one species according to equation (4.1), and a predictive value for man determined.

D.BRUCE CAMPBELL 47

Figure 4.1. Allometric scaling of Fotemustine clearance compared with the body weight in various species.

Body surface area (BSA) It has been suggested that the body surface area provides a good measure of overall metabolic rate and that this may be a better measure of relative clearance between species (Chiou and Hsu, 1988). The BSA has therefore been cal culated for each species using Meehs Formula, BSA=0.103·W0.67 (Spector, 1956) and the ratio of human BSA to animal BSA multiplied by the animal clearance, to determine the predicted human clearance. Life span correction For some drugs, particularly those which are extensively metabolised but have a low hepatic clearance, such as phenytoin, antipyrine or caffeine (Boxenbaum, 1982b; Bonati et al., 1984–5), these simple scaling methods seem to be poorly predictive for man and an allometric correction using maximum life potential (MLP) has been used to improve the accuracy (Figure 4.2). Although the allometric approach using body weight alone is

48 SIZING UP THE PROBLEM OF EXPOSURE EXTRAPOLATION

Figure 4.2. Comparison of the allometric interspecies scaling for phencyclidine using: (top) clearance (CL), and (bottom) clearance corrected for maximum life potential (MLP) in seven species (redrawn from Owens et al., 1987).

valid for many physiological functions it is poorly predictive of longevity or maximum life potential in man. Using a derived equation based on body weight alone, humans should only live for 26.6 years, clearly an underestimate. In fact Sacher (1959) has shown that a better measurement of life span can be calculated using not only body weight but also brain weight (equation (4.3)), and with this correction the MLP for man becomes 113 years (Boxenbaum and De Souza, 1988).

D.BRUCE CAMPBELL 49

(4.3) Simplistically it has been suggested that these differences in longevity can be explained by the assumption that in any one species there is a predetermined or fixed amount of total ‘body metabolic potential’ and once this is used up the animal dies (Boddington, 1978). Boxenbaum (1986) has extrapolated this concept to include intrinsic hepatic metabolism suggesting that there is a certain quantity of ‘hepatic pharmacokinetic stuff’ per unit of body weight available in a life-time which can be interrelated by the formula: (4.4) where CL is the unbound clearance, and c is a constant for each compound. Thus, the longer the animal lives, the slower this ‘stuff’ is used up. Examination of the data available from 13 disparate compounds (Table 4.1), where at least four species have been investigated, shows the MLP correction has produced good results with an exponent b equal to unity. Thus this would suggest that the relative clearance between species is directly proportional to their body weight (W) and MLP, and that animal (CL(A)) and human clearance (CL(H)) can be simply related according to equation (4.4). (4.5) The maximum life potential (MLP) has been calculated for each animal from Sacher’s formula (equation (4.3)) (mouse=2.7 y, rat=4.7 y, dog=20 y, rabbit=8 y, monkey=22 y and human=113 y). For each drug where the appropriate information was available, the human clearance has been calculated from each species using the above approaches and compared with that observed (Table 4.2), and the percentage prediction measured as:

Results The data from 60 different compounds were used in this ongoing analysis and as could be expected more data were available for the rat (n=47) compared to mouse (n=27) and dog (n=28), rabbit (n=24), or monkeys (n=17). In four cases, valproic acid, diazepam, ceftizoxime and theophylline, different results were found and data have been analysed separately. For two classes of drugs, β-lactams and benzodiazepines, data from a number of compounds were available (n=6 and 12, respectively), but only mean values were used in this analysis to minimise a class of compounds bias in the results.

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Table 4.1 Comparison of exponential values for b with MLP corrected clearance (CLu·MLP=aWb)

From Figure 4.3 it can be seen that for most species the use of the simple exponent 0.7 provided the worst prediction, particularly in the mouse and dog, which overestimated the human clearance by approximately 600 and 400 per cent, respectively. The rat and rabbit (100–150 per cent) were better but the monkey was best giving a small overestimate (36 per cent). The body surface area calculation for most animals gave a better result particularly for the rat (48 per cent) and monkey (−28 per cent), but the best method overall is the use of the maximum life potential correction which provided reasonable predictions, within 50 per cent, for all species with the exception of the mouse (89 per cent). The mean accuracy values only provide part of the picture on predictions and the variation, range and outliers can give additional information on precision and confidence of the analyses. Table 4.3 shows that although there is reasonable accuracy with the rat, rabbit and dog, the coefficients of variations and range of values for these species are large, particularly in the dog, even though the mean value is reasonable. However for the monkey most estimates of human clearance fall within close proximity to the mean provid ing good confidence in the data. Similarly the number of all compounds which have a predictability of more than 100 per cent error was large for the dog (18 per cent) and mouse (11 per cent), less for the rat and rabbit, but none were found for the monkey. In the rat, where the largest number of compounds were examined (n=56), there is a good correlation (r=0.81, p<0.01) between predicted and observed clearances using the MLP correction (Figure 4.4), but where there are inaccuracies in prediction these mostly occur for those chemicals with lower clearances.

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Figure 4.3. Mean prediction values (percentage error) for human clearance calculated for various species using: exponent 0.7, body surface area (BSA), and maximum life potential correction (MLP).

For these life span corrections, equation (4.3) has been used to calculate MLP, but monkeys in captivity, in contract organisations and zoos (Carmac, 1994), appear to live longer than the calculated 22 years and ages of 35 years are not uncommon. Substituting this longer life span into the clearance MLP correction improves the mean accuracy to −14 per cent, but the range increases and 2 per cent of compounds now give a prediction greater than 100 per cent. Attempts to combine predictions from two or more animals did not improve the accuracy of the predictions but did marginally improve the confidence of these values, particularly when the data from rat and monkey were averaged, from a confidence interval of ±20 and ±23 for rat and monkey respectively, when used alone, to ±15 when the results were combined. From this analysis of the data it would appear that measurement of the clearance of a drug in the monkey together with a correction for MLP differences, provide the best overall estimate of human clearance with the greatest confidence in the results, although for many compounds the rat or even the rabbit are good alternatives. The mouse and the dog, on the other hand, seem to be poorer animal models to extrapolate to human kinetics. Discussion There has in the past been a hesitation to use allometric scaling to predict the clearance in man, but it would appear from this review of the literature that this approach can be used for predictive purposes with an acceptable degree of accuracy, even when the clearance is measured in only one species. To put this in perspective, if the actual human clearance was 500 ml min−1, the predicted clearance using rat or monkey with MLP correction would be approximately 300 ml min −1 with a 95 per cent confidence,

Table 4.2 Human unbound clearances of the compounds used in this analysis

52 SIZING UP THE PROBLEM OF EXPOSURE EXTRAPOLATION

b

a

Campbell DB, 1993 unpublished data. CL=812 ml min−1

D.BRUCE CAMPBELL 53

54 SIZING UP THE PROBLEM OF EXPOSURE EXTRAPOLATION

Table 4.3 Interspecies comparisons of human clearance predictions expressed as percentage from observed clearances using maximum life potential corrections

a

MLP=22 years. years. c Percentage of compounds with a predicted human clearance greater than 100% of that observed. b MLP=35

Figure 4.4. Relationship between observed human clearance and that calculated from the rat using a maximum life potential (MLP) correction (n=56) (—— line of identity).

ranging from 240 to 360 ml min−1. The monkey appears to be slightly better than the rat and rabbit in terms of the accuracy of prediction, and with a few exceptions may be phylogenetically more acceptable. This is perhaps not surprising since most studies which have examined species differences in metabolism indicate that the monkey is more similar to man compared to the rat (Caldwell, 1981). All the primate data reported, as far as can be ascertained, have come from the Rhesus or Cynomolgus, Old World Macaque monkeys. The same considerations may not be true for New World monkeys, such as the squirrel or marmoset, but few kinetic comparisons have been made with these species. In practice, prediction of human clearance would involve measuring the intravenous or intramuscular kinetics, namely the infinite area under the curve, for each investigatory compound in two to four animals, together with an estimate of the in vitro protein binding in the animal under investigation and in human plasma, to obtain the free intrinsic clearance

D.BRUCE CAMPBELL 55

and then multiply the animal clearance by the ratio of weight and MLP, approximately 13 for the rat and 3.5 for a macaque monkey. Of course, as shown by these data, there can be exceptions, and the monkey and indeed the rat may not be a suitable species to undertake allometric scaling for all compounds. However there is an increasing use of in vitro systems such as isolated microsomes, hepatocytes or hepatic slices, to compare the metabolic profiles of compounds in animals. If undertaken in conjunction with allometric scaling, profound interspecies differences in the rates and extent of metabolism compared to humans could be observed and provide information on which is the most suitable species to use for scaling. Since the allometric scaling for volume appears for most compounds to be directly proportional to body weight with an exponent of approximately 1. 0, half-life can also be easily calculated thereby providing all the necessary kinetic parameters to simulate plasma levels after repeated dosing in man. With this information the absolute need to undertake kinetic analysis of industrial chemicals in volunteers would be reduced since the exposure calculated by this procedure is considerably better than that employed presently using uncertainty factors, giving errors in excess of 1000 per cent. Further studies are of course needed to confirm these initial observations, particularly with those chemicals used in industry or potential environmental pollutants, but perhaps this re-evaluation shows that allometry, when correctly used, may well have a practical role in the evaluation of their potential risk to man. References ADOLPH, E.F., 1949, Quantitative relations in the physiological constituents of mammals, Science, 109, 579–85. ANDERSEN, M.E., CLEWELL, H.J.III, GARGAS, M.L., SMITH, F.A. and REITZ, R.H., 1984, Physiologically-based pharmacokinetics and the risk assessment process for methylene chloride, Toxicol. Appl. Pharmacol., 87, 185–205. BÄÄRNHIELM, C., DAHLBÄCK, H. and SKǺNBERG, I., 1986, In vivo pharmacokinetics of felodipine predicted from in vitro studies in rat, dog and man, Acta Pharmacol. Toxicol, 59, 113–22. BACHMANN, K., 1989, Predicting toxicokinetic parameters in humans from toxicokinetic data acquired from three small mammalian species, J. Appl. Toxicol., 9(5), 331–8. BODDINGTON, M.J., 1978, An absolute metabolic scope for activity, J. Theor. Biol., 75, 443–9. BONATI, M., LATINI, R., TOGNONI, G., YOUNG, J.F. and GARATTINI, S., 1984–5, Interspecies comparison of in vivo caffeine pharmacokinetics in man, monkey, rabbit, rat and mouse, Drug Metab. Rev., 15(7), 1355–83.

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BOXENBAUM, H., 1982a, Comparative pharmacokinetics of benzodiazepines in dog and man, J. Pharmacokin. Biopharm., 10, 417–26. BOXENBAUM, H., 1982b, Interspecies scaling, allometry, physiological time, and the ground plan of pharmacokinetics, J. Pharmacokin. Biopharm., 10(2), 201– 27. BOXENBAUM, H., 1984, Interspecies pharmacokinetic scaling and the evolutionarycomparative paradigm, Drug Metab. Rev., 15(5, 6), 1071–121. BOXENBAUM, H., 1986, Time concepts in physics, biology and pharmacokinetics, J. Pharm. Sci., 75(11), 1053–62. BOXENBAUM, H. and FERTIG, J.B., 1984, Scaling of antipyrine intrinsic clearance of unbound drug in 15 mammalian species, Europ. J. Drug Metab. Pharmacokin., 9(2), 177–83. BOXENBAUM, H. and DE SOUZA, R., 1988, Physiological models, allometry, neoteny, space time and pharmacokinetics, in Peale, A. and Resagno, A. (eds) Pharmacokinetics: Mathematical and Statistical Approaches, New York: Plenum Publishing. CALABRESE, E.J., 1983, Principles of Animal Extrapolation, New York: John Wiley. CALDWELL, J., 1981, The current status of attempts to predict species differences in drug metabolism, Drug Metab. Rev., 12, 221–37. CAMPBELL, D.B. and INGS, R.M.J., 1988, New approaches to the use of pharmacokinetics in toxicology and drug development, Human Toxicol, 7, 469–79. CARMAC, M., 1994, Personal communication, London Zoo: Head Keeper of Primates. CHAPPELL, W.R. and MORDENTI, J., 1991, Extrapolation of toxicological and pharmacological data from animals to humans, Adv. Drug Res., 20, 1–116. CHIOU, W.L. and Hsu, F-H. 1988, Correlation of unbound plasma clearances of fifteen extensively metabolized drugs between humans and rats, Pharm. Res., 5 (10), 668–72. CHUNG, M., RADWANSKI, E., LOEBENBERG, D., LIN, C-C., ODEN, E., SYMCHOWICZ, S., GURAL, R.P. and MILLER, G.H., 1985, Interspecies pharmacokinetic scaling of Sch 34343, J. Antimicrob. Chemother., 15 (Supp.C), 227–33. CRUMP, K.S., 1984, A new method for determining available daily intakes, Fundam. Appl. Toxicol., 4, 854–71. DEDRICK, R.L., 1973, Animal scale-up, J. Pharmacokin. Biopharm., 1(5), 435–61. DOURSON, M.L. and STARA, J.F., 1983, Regulatory history and experimental support of uncertainty (safety) factors, Regul. Toxicol. Pharmacol, 3, 224–38. DUTHU, G.S., 1985, Interspecies correlation of the pharmacokinetics of erythromycin, oleandomycin and tylosin. J. Pharm. Sci., 74(9), 943–6. GASPARI, P. and BONATI, M., 1990, Interspecies metabolism and pharmacokinetic scaling of theophylline disposition, Drug Metab. Rev., 22(2, 3), 179–207. HINDERLING, P.H., DILEA, C., KOZIOL, T. and MILLINGTON, G., 1993, Comparative kinetics of sematilide in four species, Drug Metab. Dispos., 21(4), 662–9.

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IBRAHIM, S.S. and BOUDINOT, F.D., 1988, Pharmacokinetics of 2′,3′dideoxycytidine in rats: application to interspecies scale-up, J. Pharm. Pharmacol, 41, 829–34. KLOTZ, U., ANTONIN, K.H. and BIECK, P.R., 1976, Pharmacokinetics and plasma binding of diazepam in man, dog, rabbit, guinea pig and rat, J. Pharm. Exper. Therap., 199(1), 67–73. KURIHARA, A., NAGANUMA, H., HISAOKA, M., TOKIWA, H. and KAWAHARA, Y., 1992, Prediction of human pharmacokinetics of panipenembetamipron, a new carbapenem, from animal data, Antimicrob. Agents Chemother., 36(9), 1810–16. LAPKA, R., REJHOLEC, V., SECHSER, T., PETERKOVA, M. and SMID, M. 1989, Interspecies pharmacokinetic scaling of metazosin, a novel alphaadrenergic antagonist, Biopharm. Drug Dispos., 10, 581–9. LEBEL, M., PAONE, R.P. and LEWIS, G.P., 1983, Effect of probenecid on the pharmacokinetics of ceftizoxime, J. Antimicrob. Chemother., 12, 147–55. LINDSTEDT, S.L. and CALDER, W.A., 1981, Body size, physiological time and longevity of homeothermic animals, Quart. Rev. Biology, 56, 1–16. McGOVERN, J.P., WILLIAMS, M.G. and STEWART, J.C., 1988, Interspecies comparison of acivicin pharmacokinetics, Drug Metab. Dispos., 16(1), 18–22. MITSUHASHI, Y., SUGIYAMA, Y., OZAWA, S., NITANAI, T., SASAHARA, K., NAKAMURA, K-I., TANAKA, M. et al. 1990, Prediction of ACNU plasma concentration-time profiles in humans by animal scale-up, Cancer Chemother. Pharmacol, 27, 20–6. MORDENTI, J., 1985, Forecasting cephalosporin and monobactam antibiotic halflives in humans from data collected in laboratory animals, Antimicrob. Agents Chemother., 27(6), 887–91. MORDENTI, J., 1986, Man versus beast: pharmacokinetics in mammals, J. Pharm. Sci., 75(11), 1028–40. MROSZCZAK, E.J., LEE, F.W., COMBS, D., SARNQUIST, F.H., HUANG, B-L., WU, A.T., TOKES, L.G. et al., 1987, Ketorolac thromethamine absorption, distribution, metabolism, excretion and pharmacokinetics in animals and humans, Drug Metab.Dispos., 15(5), 618–26. MURAKAWA, T., SAKAMOTO, H., FUKADA, S., NAKAMOTO, S., HIROSE, T., ITOH, N. and NISHIDA, M., 1980, Pharmacokinetics of ceftizoxime in animals after parenteral dosing, Antimicrob. Agents Chemother., 17(2), 157– 64. OWENS, S.M., HARDWICK, W.C. and BLACKALL, D., 1987, Phencyclidine pharmacokinetic scaling among species, J. Pharm. Exper. Therap., 242(1), 96– 101. PAXTON, J.W., KIM, S.N. and WHITFIELD, L.R., 1990, Pharmacokinetic and toxicity scaling of the antitumor agents amsacrine and CI-921, a new analogue, in mice, rats, rabbits, dogs and humans, Cancer Res., 50, 2692–7. PETERS, H.P., 1983, The Ecological Implications of Body Size, Cambridge: Cambridge University Press. PUIGDEMONT, A., GUITART, R., DE MORA, F. and ARBOIX, M., 1991, Prediction of the disposition of propafenone in humans and dogs from pharmacokinetic parameters in other animal species, J. Pharm. Sci., 80(12), 1106–9.

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REIGNER, B.C., BOIS, F.Y. and TOZER, T.N., 1993, Pentachlorophenol carcinogenicity: extrapolation of risk from mice to humans, Human Exp. Toxicol., 12, 215–25. RITSCHEL, W.A., VACHHARAJANI, N.N., JOHNSON, R.D. and HUSSAIN, A. S., 1991, Interspecies scaling of the pharmacokinetic parameters of coumarin among six different mammalian species, Meth. Find Exper. Clin. Pharmacol, 13 (10), 697–702. SACHER, G.A., 1959, Relation of lifespan to brain weight and body weight in mammals, in Wolstenholme, G.E.W. and O’Connor, M. (Eds), The Life Span of Animals, pp. 115–41. Boston, Mass: Little Brown. SAWADA, Y., HANANO, M., SUGIYAMA, Y. and IGA, T., 1984, Prediction of the disposition of β-lactam antibiotics in humans from pharmacokinetic parameters in animals, J. Pharmacokin. Biopharm., 12(3), 241–61. SAWADA, Y., HANANO, M., SUGIYAMA, Y. and IGA, T., 1985, Prediction of the disposition of nine weakly acidic and six weakly basic drugs in humans from pharmacokinetic parameters in rats, J. Pharmacokin. Biopharm., 13(5), 477–92. SPECTOR, W.S., 1956, Handbook of Biological Data, Philadelphia: W.B.Saunders. SWABB, E.A. and BONNER, D.P., 1983, Prediction of aztreonam pharmacokinetics in humans based on data from animals, J. Pharmacokin. Biopharm., 11(3), 215–23. TSUNEKAWA, Y., HASEGAWA, T., NADAI, M., TAKAGI, K. and NABESHIMA, T., 1992, Interspecies differences and scaling for the pharmacokinetics of xanthine derivatives, J. Pharm. Pharmacol, 44, 594–9.

PART TWO Reactive industrial chemicals

5 Metabolism of Reactive Chemicals PETER J.van BLADEREN1,2 and BEN van OMMEN1 1 2

TNO Toxicology, Zeist

Agricultural University, Wageningen

Introduction For the purpose of the present paper, a reactive chemical will be defined as a strongly electrophilic agent. Such compounds can bind to the numerous macromolecular targets in the cell, and thus elicit toxic effects. Binding to DNA can result in mutations or cancer, binding to proteins or membrane components to cytotoxicity or specific forms of toxicity. A scale could be drawn up for the reactivity of electrophiles. However, it is not certain that those compounds on the high end of the scale, i.e. the most reactive, would also be the most toxic. On the contrary, these compounds might be expected to react quickly with water and thus not reach their target molecules. For the purpose of classifying the reactivity of electrophiles, the most useful is the theory of soft and hard acids and bases (e.g. Commandeur and Vermeulen, 1990). In principle, the preferential targets for electrophiles can be derived. Furthermore, an electrophile showing the highest affinity for the relatively hard nitrogen and oxygen nucleophiles of DNA may pose a higher risk for mutations and cancer than one reacting preferentially with soft sulfur nucleophiles such as found in proteins and glutathione. The following classes of electrophiles will be discussed: (1) quinones, which can both arylate as well as cause toxicity through redox cycling; (2) derivatives with an actual leaving group such as methylene chloride and ethylene dibromide, and (3) reagents such as isothiocyanates, isocyanates and α,β-unsatu-rated ketones and aldehydes. The enzymes involved in activation and detoxication To become toxic, almost all of the chemicals to which man is exposed, including the carcinogens, need metabolic activation. The reactive intermediates that are formed during metabolism are responsible for binding to cellular macromolecules which very likely elicit the toxic

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response. In general, other biotransformation enzymes can detoxify these metabolites. Thus, the concentration of the ultimate carcinogen, or toxicant in general, is the result of a delicate balance between the rate of activation and the rate of detoxification. Although toxicological processes can be much more complex, interindividual differences in susceptibility are certainly also a result of interindividual differences in this balance between metabolic activation and detoxification. The enzymes which are to a large extent responsible for the formation of reactive metabolites belong to the family of cytochromes P-450. However, for almost all enzymes involved in biotransformation, examples have been described of activation of specific classes of chemicals. The main classes of enzymes involved in detoxifying chemicals which are reactive per se as well as reactive metabolites are the epoxide hydrolases and the glutathione Stransferases. NADPH quinone reductase is involved in the reduction of quinones. Epoxide hydrolases Metabolites which contain an epoxide moiety may undergo hydrolytic cleavage to less reactive vicinal dihydrodiols. This reaction is catalyzed by the enzyme epoxide hydrolase (EH), which was first thought to be exclusively located in the endoplasmic reticulum (microsomal epoxide hydrolase, mEH; Oesch, 1972). In later studies on the mammalian metabolism of certain alkyl epoxides, the existence of a cytosolic EH (cEH) was demonstrated (Gill et al., 1974). The two forms of EH have complementary substrate specificity, in that many epoxides, e.g. arene oxides, which are good substrates for mEH are poor substrates for cEH, and vice versa, e.g. trans-disubstituted oxiranes are good substrates for cEH but not for mEH (Hammock and Hasagawa, 1983). Other studies have pointed to the fact that the common nomenclature of ‘microsomal’ and ‘cytosolic’ epoxide hydrolase is not semantically precise: metabolic and immunochemical studies demonstrated the existence of membrane-bound forms of cEH (Guenthner and Oesch, 1983), whereas mEH-like activity was detected in cytosolic fractions of human tissue (Schladt et al., 1988). Glutathione S-transferases Glutathione is involved in a variety of vital cellular reactions. First, a large number of the various classes of xenobiotics to which man is exposed— industrial, therapeutic as well as naturally occurring chemicals—are metabolized in vivo to reactive intermediates. Such electrophilic metabolites may bind to cellular macromolecules and thus cause toxicity. The formation of glutathione conjugates, both by spontaneous reaction between the reactive species and glutathione as well as catalyzed by the

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glutathione S-transferases, is the main detoxification mechanism for electrophiles in mammalian cells (Chasseaud, 1979). Secondly, via glutathione peroxidase and the glutathione S-transferases, hydrogen peroxide and organic peroxides are detoxified, yielding glutathione disulfide as one of the products (Prohaska, 1980). Thirdly, glutathione and the glutathione S-transferases play a role in the biosynthesis of such important endogenous compounds as prostaglandins and leukotriene C4 (Söderstrom et al., 1985; Ujihara et al., 1988). In fact, in the latter case one may argue that an endogenous compound is activated by conjugation with glutathione, since leukotriene C4 is a mediator of the adverse reactions associated with asthmatic attacks (Samuelson, 1988). The GSTs are a family of isoenzymes with broad and overlapping substrate selectivity. Although membrane-bound forms of GST have been detected (Morgenstern et al., 1988), GST activity is mainly located in the cytosol. GSTs are dimers of subunits and within a dimer, each subunit functions independently of the other (Mannervik and Jensson, 1982). The GSTs are now known to be a multi-gene family of isoenzymes, which can be divided into four classes (alpha, mu, pi and theta), based on similarity in structural, physical and catalytic properties of their subunits (Ketterer and Mulder, 1990; Vos and Van Bladeren, 1990). In addition to their crucial role in catalyzing glutathione conjugation, GSTs may also be important in intracellular binding and/or transport of endogenous and xenobiotic nonsubstrate ligands (Listowsky et al., 1988). The glutathione conjugates initially formed from electrophilic species are further processed via -glutamyltranspeptidase which splits off the glutamate residue, and dipeptidases which remove the glycine moiety. The resultant cysteine S-conjugates are then acetylated to give so-called mercapturic acids which are excreted into the urine (Jakoby, 1980). Interestingly, mercapturic acids were the first metabolites derived from xenobiotics to be recognized as such (Baumann and Preusse, 1879). In recent years it has become increasingly evident that glutathione conjugation is also involved in the formation of toxic metabolites from a variety of chemicals (Monks et al., 1990b). These metabolites display a wide spectrum of toxic effects, ranging from cytotoxicity to genotoxicity. The various mechanisms elucidated for the toxic action of the conjugates can be grouped as follows: (1) directly toxic glutathione conjugates may be formed from vicinal and geminal dihaloalkanes, via the formation of sulfur halfmustards; (2) from several types of glutathione conjugates active metabolites may be formed by further metabolic steps: conjugates of hydroquinones can be oxidized to give reactive quinones, and conjugates derived from haloalkenes are transformed into electrophilic species by the action of cysteine conjugate β-lyase. For both hydroquinones and haloalkenes the selective nephrotoxicity observed is the result of the targeting of the conjugates to the kidneys; (3) glutathione conjugates may

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serve as transporting and targeting agents for compounds that react reversibly with gluathione such as isothiocyanates, isocyanates and α, βunsaturated ketones (Van Bladeren, 1988). Glutathione S-transferase polymorphism Genetic variation in the expression of GST isoenzymes has been studied almost solely in man. Considerable variation, possibly indicating a polymorphism, has been observed for the human liver alpha class isoenzymes. The ratio of GSTA1 and GSTA2 subunits, as determined by HPLC, was found to range from 0.5 to over 10 (Van Ommen et al., 1990). However, a division into two groups, with average ratios of 1.6±0.3 and 3. 8±0.6 could be made, suggesting an alpha class polymorphism. In view of the fact that subunits GSTA1 and GSTA2 together make up a major portion of the GST protein in human liver this potential polymorphism merits further attention. For class mu isoenzymes a clear polymorphism has been observed in humans: iso-enzyme GSTM1a-1a was found to be expressed in only 60% of the samples analyzed (Board, 1981). In this study no account was taken of the fact that a second mu class isoenzyme, isoenzyme GSTM1b-1b was also suggested to play a part in this polymorphism. In a study on the excretion of the mercapturate derived from 1,3-dichloropropene in exposed workers, however, no difference was observed between mu-positive and mu negative subjects (Vos et al., 1991). Quinones and their glutathione conjugates Two modes of reactivity can form the basis of the toxicity associated with quinones: (i) their ability to undergo ‘redox cycling’ and to thereby create an oxidative stress (Kulkarni et al., 1978), and (ii) their electrophilicity allowing them to react directly with cellular nucleophiles such as protein and non-protein sulfhydryls (Dierickx, 1983). Since glutathione is the major non-protein sulfhydryl present in cells, it comes as no surprise that it is intimately involved in the biological effects of quinones. On the one hand, glutathione can act as a reducing agent, detoxifying quinones by converting them to hydroquinones with the concomitant formation of glutathione disulfide. On the other hand quinone and hydroquinonethioethers are formed. Recently considerable evidence has been gathered, indicating that a variety of these thioethers possess biological activity (Dierickx, 1983; Koga et al., 1986). The target sites for the biological (toxicological) activity of quinonethioethers is to a large extent determined by the glutathione moiety: as will be discussed, the main targets are the kidney (Monks et al., 1985) and various enzymes using glutathione as a (second) substrate, e.g. the glutathione S-transferases (Van Ommen et al., 1988). Bromobenzene is

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toxic to proximal renal tubules. The nephrotoxic effect of o-bromophenol and bromo hydroquinone was found to be considerably higher, indicating that these compounds were situated along the main bioactivation route (Monks et al., 1985). Subsequent elegant work by Monks and Lau has shown that in fact the nephrotoxicity is caused by the glutathione derivatives of bromohydroquinone (Lau and Monks, 1990). Interestingly, the relative toxicity of the quinoneglutathione conjugates increases as the extent of glutathione addition increases, i.e. the diglutathionyl derivative is more toxic than the monoconjugate (Monks et al., 1988b). The tissue selectivity is a consequence of their targeting to renal proximal tubule cells by the brushborder -glutamyl transpeptidase. AT-125, a selective inhibitor of this enzyme in vivo, protects the kidney from the toxic effects of the conjugates. The toxicity of these hydroquinone conjugates is apparently not mediated by cysteine conjugate β-lyase catalyzed formation of thiols. The inhibitor of the lyase, amino-oxyacetic acid, had only minor effects on the extent of toxicity, and the putative product, 6-bromo-2,5dihydroxythiophenol, needed activation by oxidation before it exerted any biological effect (Monks et al., 1990b). Thus, the effects of these conjugates apparently are a consequence of their oxidation to the corresponding quinones. Several isomers of 2-bromo-glutathionyl as well as the bromodiglutathionyl hydroquinones were isolated and tested. Instead of a direct correlation of toxicity with the electrochemical properties of these compounds, it was found that the diglutathionyl derivative, which is by far the most toxic, was the most stable to oxidation at pH 7.4 (Monks and Lau, 1990). The paradox was clarified by Monks and Lau by determining the oxidation potentials of the breakdown products for the mercapturic acid pathway: hydrolysis of the glutathione moiety gives rise to the cysteine derivative, which is more readily oxidized than the parent compound (Monks and Lau, 1990). Apparently two detoxication pathways are possible for these cysteine derivatives: N-acetylation results in formation of the mercapturic acid which again is relatively resistant to oxidation, but oxidative cyclization of cysteinylglycine and cysteine derivatives has been found to give 1,4-benzothiazines, which do not possess any apparent toxic properties (Monks and Lau, 1990). The action of -glutamyl transpeptidase can thus result in both activation as found for 2bromohydroquinone derivatives, but also in detoxication as was observed for 2,5-dichloro-3-(glutathion-S-yl)hydroquinone and 2,5,6-trichloro-3glutathion-S-yl)hydroquinone (Mertens et al., 1991). The ease with which the 1,4-benzothiazines are formed is very likely the determining factor in this case. A similar pathway has been worked out for p-aminophenol, a known nephrotoxic metabolite of acetaminophen (Eckert et al., 1989, 1990). Bioactivation of halogenated benzenes has long been thought to be the result of oxidation to an epoxide. However, recent studies have shown that

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the covalent binding to cellular macromolecules is not only the result of the first oxidative step, but also of the second, the formation of a quinone or hydroquinone from the initially formed phenol. The quinone in turn can be detoxified by glutathione conjugation. However, although glutathione protects the liver against toxicity due to these quinones, the conjugates are transported to the kidney and are there activated to new reactive intermediates. Thus, increasing the relative amount of glutathione Stransferases in this case would not really protect the organism, but merely change the target organ of the active metabolites. Chemicals with a leaving group Methylene chloride Both vicinal and geminal haloalkanes are bioactivated via conjugation with glutathione. The glutathione-dependent metabolism of the important industrial solvent dichloromethane yields S-chloromethyl-glutathione as the initial metabolite (Ahmed and Anders, 1976). This intermediate is held responsible for the carcinogenicity of dichloromethane in the mouse. Interestingly, this compound does not cause tumors in rats, and this has been related to the fact that the rate of metabolism via the glutathione pathway, catalyzed by the glutathione S-transferases, is much lower in rat tissue than in mouse tissue. Man has been postulated to resemble the rat in this respect and is thus presumably safe from the carcinogenic effects of methylene chloride (ECETOC, 1988). When it does not react with cellular macromolecules, the intermediate S-chloromethyl-glutathione is converted non-enzymatically to S-hydroxymethyl-glutathione, which easily eliminates formaldehyde and regenerates glutathione (Ahmed and Anders, 1978). The glutathione S-transferase isoenzyme involved in the formation of Schloromethylglutathione belongs to class theta. Interestingly, a considerable amount of interindividual variation could be observed in a group of 22 individuals (Bogaards et al., 1993). 1,2-Dibromoethane and 1,2-dichloroethane The vicinal dihaloalkanes are exemplified by 1,2-dibromoethane and 1,2dichloroethane, which are mutagenic, carcinogenic as well as nephrotoxic (Van Bladeren et al., 1980; Wong et al. 1982; Guengerich et al., 1984; Elfarra and Anders, 1985; Cheever et al., 1990). The metabolism of these compounds involves two pathways, cytochrome P-450 dependent oxidation and glutathione S-transferase catalyzed formation of glutathione conjugates. The oxidative pathway results in chloro- and bromoacetaldehyde, respectively. These aldehydes are electrophilic and

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thought to be responsible for the covalent binding of 1,2-dichloro- and 1,2dibromoethane metabolites to protein (Inskeep and Guengerich, 1984). Glutathione conjugation results in the formation of S-2haloethylglutathione derivatives, which are sulfur halfmustards and as such highly reactive metabolites (Jean and Reed, 1989). The formation of these conjugates is catalyzed by the glutathione S-transferases, and both in rat and man the alpha-class isoenzymes have been found to be the most efficient in this catalysis (Cmarik et al., 1990). The glutathione pathway is responsible for the mutagenicity (Van Bladeren et al., 1980), the DNA-binding (Koga et al., 1986) as well as very likely the carcinogenicity (Cheever et al., 1990) of 1,2-dichloro- and 1,2dibromoethane. The S-2-haloethylglutathione derivatives are strong alkylating agents (e.g. Jean and Reed, 1989). Their electrophilicity is attributable to neighboring-group assistance. The halogen atom is displaced by the sulfur atom on the next carbon atom, to form a highly reactive episulfonium ion. The intermediacy of this reactive species is supported by stereochemical studies as well as NMR data (Van Bladeren et al., 1979; Dohn and Casida, 1987; Peterson et al., 1988). The relative importance of the oxidative and glutathione-dependent pathway in vivo is difficult to determine, since both pathways give rise to the formation of the same 2-hydroxyethylmercapturate. Using tetradeutero-1,2-dibromoethane, the ratio of the pathways has been calculated as 4:1 (Van Bladeren et al., 1981b). However, isotope effects might have a considerable influence on this ratio (White et al., 1983). The major DNA-adduct derived from 1,2-dibromoethane has been identified by Guengerich and coworkers to be S-(2-(N7-guanyl)-ethyl) glutathione (Ozawa and Guengerich, 1983; Koga et al., 1986). In addition, the structure of one of several minor adducts was recently found to be S-(2(Nl-adenyl)ethyl) glutathione (Dong-Hyun et al., 1990). A series of S-2haloethylglutathione and -cysteine derivatives has been synthesized: all were found to react with DNA, specifically with guanine residues. As expected for a mechanism known to involve an intermediate episulfonium ion, adduct levels were similar for chloro- and bromo-substituted derivatives. However, in Salmonella typhimurium TA100 a large variation was observed in the ratio of mutations of adducts, indicating that the structure of the adduct has a major influence on the mutagenicity (Humphreys et al., 1990). Not all vicinal dihaloalkanes seem to give rise to the formation of episulfonium ions. Methyl substitution for instance effectively hinders the mutagenicity through this pathway (Van Bladeren, et al., 1981a) and studies on 1,2-dibromopropane (Zoetemelk et al., 1986) and hexadeuterol,2-dichloro-propane (Bartels and Timchalk, 1990) indicate that the resulting mercapturic acids are only formed through an oxidative pathway. However, for the heavily used agricultural chemical l,2-dibromo-3-

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chloropropane evidence has been accumulating recently, implicating a glutathione-mediated activation pathway in the renal and testicular toxicity associated with this compound (Pearson et al., 1990). Interestingly, consecutive formation of two episulfonium ions can occur, and in fact bisDNA-adducts have been identified (Humphreys et al., 1991). 1,2Dibromochloropropane could thus cross-link DNA strands as the initial step leading to cell death. Isoenzyme selectivity for both primary reactions has been studied extensively. The alpha and theta class glutathione S-transferases are responsible for the conjugation of EDB both in rats and man. For both of these enzymes enormous differences in levels between individuals have been found, which may be due to genetic differences, but are certainly also influenced by induction. One might expect individuals with an increased relative amount of glutathione S-transferases to be at increased risk. Reversible glutathione conjugates acting as transporting agents Numerous substrates for glutathione conjugation exist where a formal addition takes place: both the glutathionyl residue and the hydrogen atom are added to the acceptor molecule. From a chemical point of view, this reaction should be relatively easily reversible. Of course, the extent of the occurrence of the reverse reaction depends on the position of the equilibrium and is influenced by such conditions as the concentration of the reactants and the pH. The biological consequences of this reaction sequence would be that the original electrophile is detoxified initially, but not permanently: it can be released again and thus appear in unexpected parts of the body. The glutathione conjugate serves as a storage or transport form for the alkylating agent. Systemic effects of highly reactive compounds might be explained in this way. For both isothiocyanates and isocyanates evidence for this pathway has been obtained. Benzyl and allyl isothiocyanate are both naturally occurring compounds that are excreted mainly as mercapturic acids in urine after administration to rats (Brüsewitz et al., 1977). However, the mercapturate in urine is unstable under basic conditions and reforms the free isothiocyanate. The glutathione, cysteine as well as N-acetyl-cysteine conjugates derived from these isothiocyanates are all toxic in vitro (Bruggeman et al., 1986, Temmink et al., 1986). In vivo, the fact that the conjugates are somewhat more unstable in urine probably plays a role in the effects. Benzyl isothiocyanate is used for the treatment of bladder infections (Brüsewitz et al., 1977), while allyl iso-thiocyanate causes bladder tumors in male rats (Dunnick et al., 1982). The extremely reactive and toxic methyl isocyanate, used in the manufacture of carbamate pesticides, was released into the atmosphere in

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large amounts during a disaster in 1984. To explain the systemic effects of exposure to this compound, Baillie and coworkers hypothesized that these are mediated by the glutathione conjugates (Pearson et al., 1990). In fact, a rapid distribution of radioactivity throughout the body was found for rats exposed to 14C-methyl isocyanate vapor (Ferguson et al., 1988), the glutathione conjugate was identified in bile (Pearson et al, 1990) and the mercapturic acid was identified as a major urinary metabolite (Slatter et al., 1991) of rats dosed with methyl isocyanate. As was found for the isothiocyanates, in aqueous solution the synthetic glutathione conjugates are in equilibrium with the free electrophiles and glutathione: when an excess of cysteine is added to the solution, the corresponding cysteine conjugate is formed rapidly (Pearson et al., 1990). It should be realized however, that although thiols are the prime targets of iso- thiocyanates and isocyanates, the reactions with oxygen and nitrogen nucleophiles also occur and give rise to adducts that are much more stable (Pearson et al., 1991). The veterinary drug furazolidone is metabolized to a reactive metabolite that possesses an α, β-unsaturated ketone functionality. A reversible, socalled Michael adduct of this metabolite with glutathione was identified and has been suggested to play a role in the toxic effects of furazolidone (Vroomen et al., 1987). In fact residues of this metabolite covalently bound to microsomal protein could be trapped by an excess of mercaptoethanol and the glutathione conjugate gives rise to covalent binding to microsomal protein (Vroomen et al., 1988). Similarly, 2-methylfuran is metabolized to acetyl acrolein. The glutathione conjugate derived from this metabolite is unstable, and in fact toxicity of 2-methylfuran is potentiated by increasing glutathione levels by the administration of the cysteine precursor L-2oxothiazolidine-4-carboxylate (Ravindranath and Boyd, 1991). Thus, the reversibility of glutathione conjugation reactions warrants further investigation. The fact that reactive intermediates can be reformed might have important implications for the explanation of effects at sites distant from the site of initial exposure and/or initial conjugation. Conclusion Reactive chemicals can be detoxified fairly efficiently by several ubiquitous biotransformation enzymes. However, numerous cases have been reported where the initial detoxification is not the end of the story. The various pathways that the initially formed metabolites may undergo can result in unexpected toxicities at sites distant from the point of entry into the body of the electrophilic xenobiotic or the site of formation of the electrophilic metabolite.

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OMMEN, B.VAN, BESTEN, C.DEN, RUTTEN, A.C.M., PLOEMEN, J.H.T.M., Vos, R.M.E., MÜLLER, M. and BLADEREN, P.J.VAN, 1988, J. Biol Chem., 263, 12939–12942. OMMEN, B.VAN, BOGAARDS, J.J.P., PETERS, W.H.M., BLAAUBOER, B. and BLADEREN, P.J.VAN, 1990, Biochem. J., 269, 609–13. OZAWA, N. and GUENGERICH, F.P., 1983, Proc. Nat. Acad. Sci. USA, 80, 5266– 70. PEARSON, P.G., SODERLUND, E.J., DYBING, E, and NELSON, S.D., 1990a, Biochemistry, 29, 4971–81. PEARSON, P.G., SLATTER, G.T., RAKSHED, M.S., MAN, D.M., GRILLO,M. P. and BAILLIE, T.A., 1990b, Biochem. Biophys. Res. Commun., 166, 245– 50. PEARSON, P.G., SLATTER, J.G., RASHED, M.S., HAN, D.H., and BAILLIE, TH.A, 1991, Chem. Res. Toxicol, 4, 436–44. PETERSON, L.A., HARRIS, T.M., GUENGERICH, F.P., 1988, J. Am. Soc., 110, 3284–91. PROHASKA, J.R., 1980, Biochim. Biophys. Acta., 611, 87–98. RAVINDRANATH, V. and BOYD, M.R. 1991, Biochem. Pharmacol., 41,(9), 1311–18. SAMUELSSON, B., 1988, Science, 220, 568–75. SCHLADT, L., THOMAS, H., HARTMANN, R. and OESCH, F., 1988, Eur. J. Biochem., 176, 715–23. SLATTER, J.G., RASHED, M.S., PEARSON, P.G., HAN, D.H, and BAILLIE, TH.A., 1991, Chem. Res. Toxicol, 4, 157–61. SMITH, M.T., EVANS, C.G., THOR, H. and ORRENIUS, S., 1985, Oxidative Stress, 91–113. SÖDERSTRÖM, M., MANNERVIK, B., ORNING, L. and HAMMARSTRÖM, S., 1985, Biochem. Biophys. Res. Commun., 128, 265–70. TEMMINK, J.H.M., BRUGGEMAN, I.M. and BLADEREN, P.J.VAN, 1986, Arch. Toxicol., 59, 103–10. UJIHARA, M., TSUCHIDA, S., SATOH, K., SATO, K. and URADE, Y. 1988, Arch. Biochem. Biophys., 264, 428–37. Vos, R.M.E., and VAN BLADEREN, P.J. 1990, Chem.-Biol. Interact. 75, 241–65. Vos, R.M.E., VAN WELIE, R.T.H., PETERS, W.M., EVELO, C.T.A., BOGAARDS, J.J.P., VERMEULEN, N.P.E. and VAN BLADEREN, P.J., 1991, Arch. Toxicol. 65, 95–9. VROOMEN, L.H.M., GROTEN, J.P., MUISWINKEL, K.VAN, VELDHUIZEN, A. VAN and BLADEREN, P.J.VAN, 1987, Chem. Biol. Interact. 64, 167–79. VROOMEN, L.H.M., BERGHMANS, M.C.J., GROTEN, J.P., KOEMAN, J.H., and BLADEREN, P.J.VAN, 1988, Toxicol. Appl. Pharmacol, 95, 53–60. WHITE, R.D., GANDOLFI, A.J., BOWDEN, G.T. and SIPES, I.G., 1983, Toxicol. Appl. Pharmacol., 69, 170–8. WITZ, G., 1989, Free Radical Biol. Med., 7, 333–49. WONG, L.C.K., WINSTON, J.M., HONG, C.B. and PLOTNICK, H., 1982, Toxicol. Appl. Pharmacol. 63, 155–65. ZOETEMELK, C.E.M., OEI, I.H., MEETEREN-WALCHI, B., ONKENHOUT, W., GEN, A.VAN DER and BREIMER, D.D., 1986, Drug Metab. Dispos. 14, 601–7.

6 Methods for the Determination of Reactive Compounds PETER SAGELSDORFF CIBA-GEIGY Ltd, Basel

Introduction It has well been recognised for a long time that adverse effects of chemicals are associated with their reactivity whereby many unreactive chemicals are metabolised in the cell to a reactive intermediate. Reactive intermediates are generally electrophiles which undergo reactions with cellular nucleophiles and the toxicological response is often the consequence of the covalent binding of a chemical to cellular macromolecules. This chapter will provide a brief survey of current methods for the determination of protein and DNA adducts generated by reactive compounds, and will discuss some useful applications of these technologies. Source of reactive metabolites Reactive chemicals are generally strong electrophilic agents. These compounds can be reactive per se (direct electrophiles), such as methylmethanesulphonate, epoxides or strained lactones. On the other hand, unreactive chemicals can be enzymatically converted to electrophilic agents (indirect electrophiles), such as aromatic amines and nitroarenes to the corresponding nitrenium ions, polycyclic aromatic hydrocarbons to diol epoxides or N-nitroso compounds to carbenium ions (Figure 6.1; Magee et al., 1975; Weissburger and Williams, 1975; Lutz, 1979). Interaction of reactive compounds with cellular constituents As electrophiles, these reactive compounds undergo reactions with nucleophiles. The nature of the toxicological response is dependent on the biological macromolecule affected. Reactions with water or glutathione, two of the most abundant cellular nucleophiles, in most cases, lead to an inactivation of the respective reactive compound.

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Figure 6.1 Examples of electrophilic compounds. The arrows indicate the suspected electrophilic centres (from Lutz, 1979).

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Figure 6.2 Nucleophilic centres in nucleobases and DNA ‘adduct library’ indicating the preferential binding sites on guanine for several classes of chemicals. The most reactive targets are indicated by an arrow (from Lutz, 1979; Beach and Gupta, 1992).

The most important nucleophilic centres in proteins are the side chains of the amino acids cysteine, methionine, histidine and tyrosine, and the amino group of the N-terminal amino acid. Reactions of electrophiles with proteins lead to the formation of protein adducts. This results in general or specific cytotoxicity depending on whether the function of a particular protein is disturbed (Lawley, 1976; Brooks, 1977). Adduct formation with blood proteins can result in the formation of immunogens and subsequent allergenic responses. Finally, reactions with DNA predominantly occur with the nucleobases adenine, cytosine, thymine and guanine, whereby the most important nucleophile in DNA is guanine (Figure 6.2). Adduct formation with nucleobases in DNA is recognised as a crucial step in the formation of mutations and cancer (Lutz, 1979).

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Methods for the determination of adducts Adduct formation of reactive compounds with protein or DNA can easily be detected by the use of radiolabelled test compounds. However, the radiolabelled compounds are often not available. In addition, real exposure situations and unknown mixtures of compounds can not be assessed. For a number of chemicals, therefore, alternative methods for adduct determinations have been developed during the past. Reactions with proteins can be assessed by analysing haemoglobin or albumin adducts. These proteins can easily be isolated in large quantities (100 mg haemoglobin, 30 mg albumin per ml blood) and with sufficient purity from the blood of treated animals or occupationally exposed humans. Methods for the determination of DNA adducts generally require a higher sensitivity, since DNA from treated animals or exposed humans is only available in small amounts (1–2 mg per g tissue, 4 µg from white blood cells per ml blood). Protein adducts Physical methods Aromatic amines and nitroarenes The key step in the metabolic activation of arylamines to the respective nitrenium ions involves N-hydroxylation. The N-hydroxylamines can be further oxidised in erythrocytes to the corresponding nitroso compounds with a concurrent production of methaemoglobin. On the other hand, nitroarenes can be metabolically reduced to the corresponding nitrosoarenes. The nitrosoarenes covalently bind to the thiol group of cysteine residues and rearrange to give stable sulphinic acid amides. Mild alkaline treatment can be used to hydrolyse these adducts. The liberated parent amines can be extracted and analysed by HPLC with specific detection methods, such as electrochemical or fluorescence detection. In order to improve the sensitivity of the assay, the extracted adducts can be derivatised with electrophores and analysed by GC with electron capture detection or by GC/MS (Bailey et al., 1990; Skipper and Tannenbaum, 1990; Sabbioni, 1992, 1994). Polycyclic aromatic hydrocarbons Polycyclic aromatic hydrocarbons are oxidised by cytochrome P450 to epoxides, which are rapidly hydrolysed. However, further oxidation to the

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Figure 6.3 Modified Edman degradation of alkylated N-terminal valine in haemoglobin (Törnqvist et al., 1986).

respective diol epoxides results in the formation of relatively stable electrophiles which also alkylate cysteine residues in proteins. Upon mild acid treatment these adducts are liberated as the respective tetrols. Similar to adducts from aromatic amines, the tetrols can be extracted and analysed by HPLC with specific detection methods, or by GC with electron capture detection or by GC/MS after derivatisation with electrophores (Shugart and Kao, 1985; Weston et al., 1989; Day et al., 1990). Alkylating agents Adducts of alkylating agents with the thiol group of cysteine, histidine or the N-terminal amino acids resist alkaline or acid hydrolysis. To determine the alkylated amino acids the protein is hydrolysed with 6 N HCl and the amino acids are separated on a anion exchange column. The fractions containing the alkylated amino acids are derivatised with electrophores and analysed by GC/MS (van Sittert et al., 1985; Bailey et al., 1987).

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Alternatively, the alkylated N-terminal valine of haemoglobin can selectively be cleaved off by a modified Edman degradation with pentafluorophenyl isothiocyanate (PFPITC). Since alkylation of the amino group favours the reaction, conditions can be selected to exclusively liberate alkylated N-terminal amino acids whilst leaving the non-adducted N-terminal valine intact (Törnqvist et al., 1986). The resulting pentafluorophenyl thiohydantoine (PFPTH) derivative can be extracted and quantified by GC/MS (Figure 6.3). Immunological methods Immunological methods have been developed for the quantification of some adducts of aromatic amines, polycyclic aromatic hydrocarbons and alkylating agents. However, these methods involve a couple of time consuming steps for the isolation of an appropriate antibody. The respective haemoglobin adduct has to be chemically synthesised, an animal has to be immunised with the modified haemoglobin and, later on, polyclonal antibodies can be isolated from the blood of the immunised animal. In order to produce monoclonal antibodies, which normally have a better specificity and sensitivity, spleen cells of the immunised animal are fused with myeloma cells and the antibodies can be isolated from the cell culture. The methods for the determination of adducts include competitive radioimmunoassays and solid phase assays (ELISA, USERIA). The protein is partially hydrolysed, adsorbed on a solid surface and treated with the primary antibody. An anti-antibody which is directed against the primary antibody, radiolabelled, or conjugated to a fluorescent dye or an indicator enzyme, is added and the amount of bound label is quantified (Santella et al., 1986; Lee and Santella, 1988). DNA adducts Physical methods Aromatic amines and nitroarenes The hydroxylamines produced by enzymatic hydroxylation of aromatic amines or by reduction of nitrosoarenes are further conjugated (Osulphatation, O-acetylation, O-glucuronidation). The conjugates can decompose to the respective nitrenium ions which add predominantly to the C8 of guanine. Similarly to protein adducts of these compounds, the adducts can be liberated from DNA by alkaline hydrolysis or hydrazinolysis, extracted and

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quantified by HPLC with fluorescence or electrochemical detection. In order to improve the sensitivity the extracted adducts can be derivatised with electrophores and analysed by GC/MS (Bakthavachalam et al., 1991; Lin et al., 1991). Polycyclic aromatic hydrocarbons The diol epoxides enzymatically produced from polycyclic aromatic hydrocarbons mainly adduct at the exocyclic amino group of guanine. The adducts can be liberated from DNA by acid hydrolysis, extracted and quantified by HPLC with fluorescence or electrochemical detection or by GC with electron capture detection or GC/MS after suitable derivatisation with electrophores (Rahn et al., 1982; Shugart and Kao, 1985; Weston et al., 1989). Alkylating agents Alkylating agents mainly alkylate the N7 of guanine but also give rise to the formation of other N- and O-alkyl nucleobase adducts. The DNA bases are liberated by hydrolysis and analysed for the presence of adducts by HPLC with electrochemical detection or they are extracted, derivatised with electrophores and analysed by GC/MS (Minnetian et al., 1987; Groot et al., 1994). Some alkylating agents and small epoxides lead to the formation of cyclic nucleobase adducts which exhibit strong fluorescence. Enzymatic or acid hydrolysis can be used for the liberation of the DNA constituents and the fluorescent adducts can be analysed by HPLC with fluorescence detection (Fedtke et al., 1990; Steiner et al., 1992a). Immunological methods Immunological methods for the determination of DNA adducts essentially follow the procedure as outlined already for protein adducts: generation of an antibody, absorption of the DNA on a solid surface, incubation with the antibody and a labelled anti-antibody. However, for the production of the antibody an additional step has to be performed. The immune system normally does not respond to small molecules. Therefore, the chemically synthesised base or nucleoside adduct has to be coupled to a carrier protein, in order to obtain an immunogen (Perera et al., 1986; Santella, 1988; Poirier, 1993). Postlabelling One of the most popular assays for determination of DNA adducts is the postlabelling assay. The DNA is enzymatically hydrolysed to the four

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natural deoxynucleoside-3′-monophosphates (dNp) and the dNp adducts. The adducted dNp carrying bulky or aromatic substituents are enriched by extraction with butanol in the presence of a phase transfer agent or by selective digestion of the natural (unadducted) dNp with nuclease P1. The enriched adducted dNp are labelled with [32P]- or [33P]ATP (Figure 6.4). Polynucleotide kinase T4 is used to catalyse the transfer of the labelled phosphate group from ATP to the 5′ position of the dNp. The labelled deoxynucleoside-3′,5′-bisphosphates are then separated by multidirectional TLC on polyethyleneimine coated cellulose. Radioactive impurities and unused ATP are running to the top with phosphate buffer (D1), whereas nucleotides carrying aromatic or bulky adducts are retained at or near the origin. The part containing the impurities and unused ATP is cut off, and the adducts are chromatographed in D3 (opposite to D1) and D4 (perpendicular to D3) with ammonia and ammonia/ propanol or urea containing buffers. Adduct spots are visualised and quantified by autoradiography and Cherenkov counting or by phosphor imaging (Gupta, 1985; Reddy and Randerath, 1986; Beach and Gupta, 1992). Alternatively, the enriched nucleotide adducts can be chemically derivatised with fluorescent labels and analysed by HPLC with fluorescence detection. However, this method does not reach the sensitivity of the radioactive assay (Sharma and Jain, 1991; Jain and Sharma, 1993). Comparison of different methods The methods used for the determination of protein and DNA adducts are summarised in Tables 6.1 and 6.2. Special attention is drawn to the cost of equipment and time required for analysis. HPLC methods with electrochemical or fluorescent detection are relatively insensitive and only applicable with compounds which are strongly fluorescent or electrochemically active. Since the costs for the equipment used and the time consumption are relatively low, these methods are attractive in certain cases. GC with electron capture detection or GC/MS offers better sensitivity. However, the method requires derivatisation. In addition, the costs for the equipment of the GC/MS methods are quite high. Immunoassays are very sensitive, but involve a number of time consuming steps for the preparation of an appropriate antibody, and are only possible if the structure of the respective adduct is known. The postlabelling method for DNA adducts offers the best sensitivity, with low equipment costs and low to medium time consumption. However, the standard method only detects bulky or aromatic adducts.

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Figure 6.4 Schematic representation of the postlabelling assay for determination of DNA adducts (Gupta, 1985; Reddy and Randerath, 1986; Beach and Gupta, 1992).

Examples/applications In the following sections some useful applications of adduct determinations, which have been performed in our laboratory, will be presented.

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Table 6.1 Methods for the determination of protein adducts

a An approximate mean sensitivity is given in pmol (=10−12 mol) adducts/g haemoglobin.

Table 6.2 Methods for the determination of DNA adducts

a An

approximate mean sensitivity is given in fmol (=10−15 mol) adducts/mg DNA.

Lack of bioavailability of 3,3′-dichlorobenzidine from diarylide pigments 3,3′-Dichlorobenzidine is an important intermediate in the production of diarylide pigments and azo dyes. Some of these pigments have been tested in long term studies and shown to exert no specific toxicological effects and to be not carcinogenic to experimental animals (ETAD Report, 1990). However, there might be a theoretical hazard after metabolic splitting of the pigments into DCB, a known animal carcinogen (IARC, 1982). DCB and its N-acetylated metabolite are N-hydroxylated and oxidised to the corresponding nitroso compound which binds to haemoglobin. Since no repair of haemoglobin adducts occurs, these adducts cumulate during the

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Figure 6.5 HPLC/ECD profiles obtained after hydrolysis and extraction of haemoglobin samples isolated from an untreated rat (control) and from rats treated for 4 weeks with DCB (2 mg kg−1), Direct Red 46 (160 mg kg−1), Pigment Yellow 13 (400 mg kg−1) and Pigment Yellow 17 (400 mg kg−l) as well as of commercially available bovine haemoglobin (Hb-bovine).

life span of the erythrocyte. Haemoglobin adduct formation, therefore, was used to monitor the liberation of DCB from diarylide pigments. Rats were treated by daily oral gavage for 4 weeks with the pigment at daily dose levels of 400 mg kg−1 body weight. As a positive control, animals were treated accordingly with DCB (2 mg kg−1) or with Direct Red 46 (160 mg kg−1), asoluble azo dye with known bioavailability of DCB. After termination of the treatment, haemoglobin was isolated and hydrolysed in 0.1 N sodium hydroxide. The liberated DCB and monoacetyl-DCB were extracted with toluene/2-propanol and analysed by HPLC with electrochemical detection. With 2 mg DCB kg−1 body weight DCB and monoacetyl-DCB adducts were clearly detectable, amounting up

P.SAGELSDORFF 83

to 50 ng g−1 haemoglobin (Figure 6.5). No macromolecular adducts were detectable in the rats treated with the two diarylide pigments. The limits of determination would correspond to a daily DCB dose of 0.3–0.5 mg kg−1 body weight, indicating that DCB was not liberated from the pigments at a determination limit of 0.3% of the DCB equivalents, whereas the bioavailability of DCB in the rats treated with the azo dye could clearly be confirmed. Formation of glycidaldehyde from glycidylethers Bisphenol A diglycidylether (BPADGE) is widely used as component of epoxy resins. The chemical reactivity of this class of compounds is a prerequisite for their technical use, and accounts for the sensitising, mutagenic and in some cases carcinogenic properties of many epoxy resin monomers. It was suggested that the metabolic inactivation of BPADGE by hydrolysis of epoxides may form an equilibrium with its metabolic activation by oxidative dealkylation of the intact glycidyl side chain followed by the release of glycidaldehyde. Cutaneous treatment of mice with glycidaldehyde led to the formation of one major epidermal DNA adduct which was identified as HMEdA (hydroxymethylethenodeoxyadenosine, Steiner et al., 1992a). This cyclic deoxyadenosine adduct is strongly fluorescent and can be quantified by fluorescence measurements. In order to investigate the formation of glycidaldehyde from BPADGE, mice were treated with BPADGE (2 mg) and the fluorescent glycidaldehydeDNA adducts formed in epidermal DNA were compared with those obtained after treatment with glycidaldehyde (2 mg). After 24– 96 h epidermal DNA was isolated, enzymatically digested to the deoxynucleoside-3'-monophosphates and analysed for the presence of HMEdA by HPLC with fluorescence detection (excitation at 231 nm, emission at 420 nm). In glycidaldehyde treated mice 166 adducts per 106 nucleotides could be detected after an exposure time of 24 h (Figure 6.6) whereas with epidermal DNA from BPADGE treated mice 0.2– 0.8 adducts per 106 nucleotides were found. This adduct level would be equal to a dose of 10 µg glycidaldehyde, indicating that, at the most, 1.1% of the glycidaldehyde moiety in BPADGE were bioavailable for DNA-adduct formation (Steiner et al., 1992b). Determination of reactive compounds in unknown mixtures A challenging task is the analysis of reactive metabolities in unknown mixtures of different compounds. In order to assess the impact of chemical pollution on aquatic organisms, rainbow trouts were continuously exposed

84 METHODS FOR THE DETERMINATION OF REACTIVE COMPOUNDS

to the diluted effluent discharges of a chemical production plant for 3 months. The plant produced different dyes and chemicals and the waste water therefore could be contaminated with a variety of aliphatic and aromatic amines and some cyclic aromatic hydrocarbons. After termination of the treatment, liver and gill DNA from exposed and control trouts was analysed by [32P]postlabelling for the presence of DNA adducts. The DNA was enzymatically hydrolysed to the nucleotides. Adducted nucleotides were extracted with butanol in the presence of the phase transfer agent tetrabutylammonium chloride and postradiolabelled with [32P]ATP and PNK. The labelled nucleotides were separated by multidirectional TLC with 1.0 M phosphate buffer, pH 6.6, in D1, 0.4 M ammonia in D3 and 4 N ammonia/propanol (1.2:1) in D4. A final development in direction D4 with 1.0 M phosphate buffer, pH 6.6, was used as background clean up. In the trouts exposed to control water no DNA adducts were detectable, neither in the livers nor in the gills (Figure 6.7). In contrast, in the trouts exposed to the highest concentration of the waste water, at least 4 DNA adducts could be found in the livers and in the gills. The overall DNA adduct level in the exposed trouts was relatively low (1 adduct per 108 nucleotides, which indicated only a minimal cancer risk for the exposed fish. Limitations However, the methods presented for adduct determination have their limitations. For protein adduct determination the most popular method is by HPLC with electrochemical or fluorescence detection after hydrolysis and extraction of the adducts. This is due to the low cost and time consumption of the method. This method is hampered by the possibility of interferences, which can elute in the range of the compounds of interest. For an exclusion of haemoglobin adducts formation at low levels it is therefore crucial to obtain additional information about the chromatographic peaks of interest, such as for example, by GC/MS. DNA adducts are often assessed by [32P]postlabelling. This method is limited by low yields of the enrichment and labelling procedures and by choosing the appropriate chromatographic conditions for the resolution of the labelled adducts. The lack of detectability of some DNA adducts, although they may contain aromatic moieties, enforces the use of a positive standard in order to check for the yield of the enrichment and the labelling reaction, and to check for appropriate chromatographic conditions to resolve the adducts.

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Figure 6.6 HPLC/fluorescence analysis of epidermal DNA hydrolysates from a control (a) and a BPADGE treated mouse (b), and UV trace of synthetic HMEdAp and HMEdGp (c).

86 METHODS FOR THE DETERMINATION OF REACTIVE COMPOUNDS

Figure 6.7 TLC chromatograms of DNA adducts in gills and livers of rainbow trouts, exposed for 3 months to waste water or control water. Top: control water, liver DNA (left chromatogram), gill DNA (right chromatogram); bottom: waste water, liver DNA (left chromatogram), gill DNA (right chromatogram).

Conclusions Each method, although inherently chemically-specific, has its advantages and limitations depending on the adduct-type. The continued rapid development of the technologies described for assessing biomarkers should result in more accurate assessment of the intracellular reactions of chemicals and thereby provide information about the mechanism of toxicity of a compound under investigation.

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Acknowledgements Grateful thanks to Drs Markus Joppich and Regula Joppich-Kuhn for haemoglobin adduct analyses and Dr Sandra Steiner for the development of the fluorescence assay for HMEdAp. References BAILEY, E., FARMER, P.B. and SHUKER, D.E.G., 1987, Estimation of exposure to alkylating carcinogens by the GC-MS determination of adducts to haemoglobin and nucleic acid bases in urine, Arch. Toxicol., 60, 187–91. BAILEY, E., BROOKS, A.G., BIRD, I., FARMER, P.B. and STREET, B., 1990, Monitoring exposure to 4,4′-methylenedianiline by the GC/MS determination of adducts to haemoglobin, Anal. Biochem., 190, 175–81. BAKTHAVACHALAM, J., ABEL-BAKY, S. and GIESE, R.W., 1991, Release of 2aminofluorene from N-(deoxyguanosine-8-yl)-2-aminofluorene by hydrazinolysis, J.Chromatogr., 538, 447–51. BEACH, A.C. and GUPTA, R.C., 1992, Human biomonitoring and the 32Ppostlabelling assay, Carcinogenesis, 13, 1053–74. BROOKS, P., 1977, The role of covalent binding in carcinogenicity, in Jollow, D.J. et al. (Eds) Biological Reactive Intermediates, pp. 470–480, New York: Plenum. DAY, B.W., NAYLOR, S., GAN, L.-S., SHALI, Y., NGUYEN, T.T., SKIPPER, P. L., WISHNOK, J.S. and TANNENBAUM, S.R., 1990, Molecular dosimetry of polycyclic aromatic hydrocarbon epoxides and diol epoxides via haemoglobin adducts, Cancer Res., 50, 4611. ETAD Report, 1990, On the carcinogenic potential of diarylide azo pigments based on 3,3′-dichlorobenzidine, T 2028-CA, Toxicological Subcommittee of ETAD (TSC). FEDTKE, N., BOUCHERON, J.A., WALKER, V.E. and SWENBERG, J.A., 1990, Vinyl chloride induced DNA adducts: II: formation and persistence of 7-(2'oxoethyl) guanine and N2,3-ethenoguanine in rat tissue DNA. Carcinogenesis, 11, 1278–1292. GREEN, L.C., SKIPPER, P.L., TURESKY, R.J., BRYANT, M.S. and TANNENBAUM, S.R., 1984, In vivo dosimetry of 4-aminobiphenyl in rats via a cysteine adduct in haemoglobin, Cancer Res., 44, 4254–9. GROOT, A.J.L., JANSEN, J.G., VAN WALKENBURG, C.F.M. and ZEELAND, A.A., 1994, Molecular dosimetry of 7-alkyl- and O6-alkylguanine in DNA by electrochemical detection, Mutat. Res., 307, 61–6. GUPTA, R.C., 1985, Enhanced sensitivity of 32P-postlabelling analysis of aromatic carcinogen: DNA adducts, Cancer Res., 45, 5656–62. IARC, 1982, Monographs of the carcinogenic risk of chemicals to humans: some industrial chemicals and dyestuffs. 3,3′-Dichlorobenzidine and its hydrochloride, International Agency for the Research on Cancer, Vol. 29, pp. 239–56. JAIN, R. and SHARMA, M., 1993, Fluorescence postlabelling assay of DNA damage induced by N-methyl-nitrosourea, Cancer Res., 53, 2771–4.

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LAWLEY, P.D., 1976, Carcinogenesis by alkylating agents, in Searle C.E. (Ed.), Chemical Carcinogens, pp. 83–244, ACS Monograph 173, Washington, DC: American Chemical Society. LEE, M.L. and SANTELLA, M., 1988, Quantitation of protein adducts as a marker of genotoxic exposure: immunologic detection of benzo[a]pyreneglobin adducts in mice, Carcinogenesis, 9, 1773–7. LIN, D.-X., LAY, J.O. JR, BRYANT, M. and KADLUBAR, F.F., 1991, Analysis of 4-aminobiphenyl-DNA adducts by alkaline hydrolysis and negative ion GC/ MS, in Biomonitoring and Susceptibility Markers in Human Cancer: Applications in Molecular Epidemiology and Risk Assessment, P-20, Haway, USA: KailuaKono. LUTZ, W.K., 1979, In vivo covalent binding of organic chemicals to DNA as a quantitative indicator in the process of chemical carcinogenesis, Mutat. Res., 65, 289–356. MAGEE, P.N., PEGG, A.E. and SWANN, P.F., 1975, Molecular mechanisms of chemical carcinogenesis, in Grundmann E. (Ed.), Handbuch der allgemeinen Pathologie, Vol. VI/6, pp. 329–419, Berlin: Springer. MINNETIAN, O., SAHA, M. and GIESE, R.W., 1987, Oxidation-elimination of a DNA base from its nucleoside to facilitate determination of alkyl chemical damage to DNA by GC/MS with electrophore detection, J.Chromatogr., 410, 453–7. PERERA, F., SANTELLA, R. and POIRIER, M., 1986, Biomonitoring of workers exposed to carcinogens: immunoassay to benzo[a]pyrene-DNA adducts as a prototype, J. Occup. Med., 28, 1117–23. POIRIER, M.C., 1993, Antisera specific for carcinogen-DNA adducts and carcinogenmodified DNA: applications for detection of xenobiotics in biological samples, Mutat. Res., 288, 31–8. RAHN, R.O., CHANG, S.S., HOLLAND, J.M. and SHUGART, L.R., 1982, A fluorimetric HPLC assay for quantitating the binding of benzo[a]pyrene metabolites to DNA, Biochem. Biophys. Res. Commun., 109, 262–8. REDDY, M.V. and RANDERATH, K., 1986, Nuclease P1-mediated enhancement of sensitivity of 32P-postlabelling test for structurally diverse DNA adducts, Carcinogenesis, 7, 1543–51. SABBIONI, G., 1992, Quantitative structure activity relationship of the Noxidation of aromatic amines, Chem.-Biol. Interact., 81, 91–117. SABBIONI, G., 1994, Haemoglobin binding of nitroarenes and quantitative structure activity relationships, Chem. Res. Toxicol, 7, 267–74. SANTELLA, R.M., 1988, Application of new techniques for the detection of carcinogen adducts to human population monitoring, Mutat. Res., 205, 271– 82. SANTELLA, R.M., LIN, C.D. and DHARMARAJA, N., 1986, Monoclonal antibodies to a benzo[a]pyrene diolepoxide modified protein, Carcinogenesis, 7, 441–4. SHARMA, M. and JAIN, R., 1991, Nuclease P1-mediated fluorescence postlabelling assay of AAF modified DNA model d(TACGTA) and calf-thymus DNA, Biochem. Biophys. Res. Commun., 177, 151–8.

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SHUGART, L. and KAO, J., 1985, Examination of adduct formation in vivo in the mouse between benzo[a]pyrene and DNA of skin and haemoglobin of red blood cells. Environm. Health Perspect., 62, 223–6. VAN SITTERT, N.J., DE JONG, G., CLARE, M.G., DAVIES, R., DEAN, B.J., WREN, L.J. and WRIGHT, A.S., 1985, Cytogenetic, immunological and haematological effects in workers in an ethylene oxide manufacturing plant, Br. J. Ind. Med., 42, 19–46. SKIPPER, P.L. and TANNENBAUM, S.R., 1990, Protein adducts in molecular dosimetry of chemical carcinogens, Carcinogenesis, 11, 507–18. STEINER, S., CRANE, A.E. and WATSON, W.P., 1992a, Molecular dosimetry of DNA adducts in C3H mice treated with glycidaldehyde, Carcinogenesis, 13, 119– 24. STEINER, S., HÖNGER, G. and SAGELSDORFF, P., 1992b, Molecular dosimetry of DNA adducts in C3H mice treated with bisphenol A diglycidylether, Carcinogenesis, 13, 969–72. TÖRNQVIST, M., MOWRER, J., JENSEN, S. and EHRENBERG, L., 1986, Monitoring of environmental cancer initiators through haemoglobin adducts by a modified Edman degradation method, Anal. Biochem., 154, 255–66. WEISSBURGER, J.H. and WILLIAMS, G.N., 1975, Metabolism of chemical carcinogens, in Becker, F.F. (Ed.), Cancer, Vol. 1, pp. 185–234, New York: Plenum. WESTON, A., BOWMAN, E.D., ROWE, M.L., MANCHESTER, D.K. and HARRIS, C.C., 1989, Fluorescence and mass spectral evidence for the formation of benzo[a]pyrene anti-diol-epoxide-DNA and -haemoglobin adducts in humans, Carcinogenesis, 10, 251–7.

PART THREE Pulmonary toxicology of industrial chemicals

7 Studies to Assess the Carcinogenic Potential of Man-Made Vitreous Fibers THOMAS W.HESTERBERG, GERALD R.CHASE, RICHARD A.VERSEN and ROBERT ANDERSON Schuller International, Inc., Littleton, CO

Introduction Man-made vitreous fibers (MMVFs) are a class of materials which have found many applications in both residential and industrial settings. MMVFs are fibrous inorganic substances that are made primarily from rock, clay, slag or glass. Sometimes referred to as man-made mineral fibers (MMMFs), the major classes of MMVF are refractory ceramic fibers (RCFs), fibrous glass, rock (stone) wool and slag wool. RCF, the smallest category of MMVF, represents only about 1–2 per cent of the world production of MMVF. It is made by melting Al2O3 and SiO2 in about equal amounts or by melting kaolin clay and then ‘spinning’ or ‘blowing’ this molten material into fibers. Most RCF is used as a high temperature furnace insulation. World production of RCF in 1990 was about 80 million 1b. Fibrous glass is the largest category of the MMVFs and is used in insulation, air handling, filtration and sound absorption. The thermal, acoustical and fire resistant properties of these products have led to their widespread use in a variety of residential and commercial applications. Production of fibrous glass in North America in 1989 was approximately 1.8 million t. Slag and rock wool are composed primarily of calcium, magnesium, aluminum and silica. Since 1975, most slag wool has been produced from the waste slag that resulted from the reduction of iron ore to iron. Rock wool fibers are made from basaltic rocks with additives such as limestone or dolomite. Slag and rock wool are used in residential and commercial low and high temperature insulation and in acoustical ceiling tiles and wall panels. About 75% of slag wool production is used in acoustical ceiling tile manufacture in North America. Animal studies and epidemiological studies have been conducted to assess the potential biological effects of MMVFs. This research has been reviewed by the International Agency for Research on Cancer (IARC, 1988), the International Programme on Chemical Safety (IPCS, 1988), and the US Environmental Protection Agency (Vu, 1988). These reviews are consistent

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in the judgement that chronic inhalation studies of airborne fibers provide the best model for assessing potential risk to man (McClellan et al., 1992). In assessing the carcinogenic risks of exposure to any possible occupational hazard, research is pursued through several different scientific techniques. Studies of mortality (analysis of death rates) are used to evaluate the potential carcinogenicity associated with direct human exposure. Animal exposure studies are used to not only evaluate the potential carcinogenicity but to also investigate the mechanisms of disease development. Industrial hygiene and engineering studies are used for quantifying exposures. Epidemiological studies By general agreement among experts (IARC, 1988; IPCS, 1988), two major historical cohort studies are considered to have comprehensively addressed the mortality experience of workers engaged in the production of FG, rock wool and slag wool: a European study conducted by the International Agency for Research on Cancer (IARC), and a University of Pittsburgh study conducted in the USA. The discussion here will concentrate on those two studies. For a summary of other studies, the reader is referred to the IARC review (IARC, 1988). There are no published reports of the mortality experience of RCF workers. Epidemiological studies of workers engaged in the manufacture of all major classes of MMVF are underway or are continuing. Morbidity studies of the respiratory health of workers are not discussed here. The IARC study IARC researchers reported their study at the WHO Occupational Health Conference on the Biological Effects of Man-Made Mineral Fibres at Copenhagen in 1982, with a follow-up in 1986 (Simonato et al., 1987). The updated study is also published in the Scandinavian Journal of Work, Environment & Health, Volume 12, Supplement 1, 1986. The mortality of 23609 workers (2836 deaths) employed in 13 European factories engaged in the production of MMVF (including 11 852 fibre glass production workers at six plants in five countries and 10 115 rock wool/slag wool production workers at seven plants in four countries) has been studied (Saracci et al., 1984) and updated (Simonato et al., 1987). The authors reported an ‘excess of lung cancer among rock-wool/slag workers employed during an early technological phase before the introduction of dust-suppressing agents’, and concluded that ‘fiber exposure, either alone or in combination with other exposures, may have contributed to the elevated risk’. The authors also reported that ‘no excess of the same magnitude was evident for glass-wool production, and the follow-up of the continuous-

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filament cohort was too short to allow for an evaluation of possible longterm effects’. It was also noted that ‘there was no evidence of an increased risk for pleural tumors or non-malignant respiratory diseases’. An update of this study is underway. The University of Pittsburgh study This study was also reported at the WHO Occupational Conference on Biological Effects of Man-Made Mineral Fibres at Copenhagen in 1982 and the follow-up Conference in 1986 (Simonato et al., 1987). Subsequent to the 1986 Conference, additional analyses were completed and included in the manuscript published for the proceedings (Enterline et al., 1987). The study has been updated and published (Marsh et al., 1990). The University of Pittsburgh researchers’ comprehensive mortality review of more than 16000 workers— many with long-term exposure up to 40 years—was undertaken at 17 US fiber-glass, rock-wool and slag-wool manufacturing plants, including 14800 fiber-glass workers in 11 plants. The original report, given in 1982, covered the mortality experience from the 1940s to the end of 1977. The same group of workers was followed through 1982 (reported in October 1986, with additional analyses available in June 1987). The June 1987 report contained, for the first time, local area mortality statistics for each of the plants as the basis for studying the mortality experience. Experts agree that, barring unusual circumstances, local area comparisons are most appropriate. The study has been further updated through 1985, with publication in 1990. For respiratory cancer, in the latest update there was a small but statistically significant increase for fiber glass production workers. However, aside from the issue of uncontrolled potential confounding, the study provides no evidence to date that respiratory cancer mortality is related to fiber glass exposure. There was a somewhat larger statistically significant excess of respiratory cancer mortality reported for slag wool and rock wool production workers. The absence of any clear exposure-response relationship for any of the fiber groups studied led the authors to conclude that ‘overall, the evidence of a relationship between exposure to man-made mineral fibers and respiratory cancer appears to be somewhat weaker than in the previous update’. Consistent with the IARC study, no increase in the occurrence of mesothelioma has been observed in this cohort. This study has now been expanded to include well over 30000 workers from 14 fiber-glass and six rock wool and slag wool facilities. Other epidemiological studies In addition to the two major studies highlighted above, a number of other studies have been conducted as well. Many of them widely overlap these

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major studies, comprise sub-groups within them, or represent smaller worker populations outside of them. A Canadian study was reported by Shannon at the WHO Occupational Health Conference on Biological Effects of Man-Made Mineral Fibers at Copenhagen in 1982 and 1986 (Shannon et al., 1987). It followed 2557 male workers at a Canadian glass wool plant through 1977 and was later updated to extend the follow-up to the end of 1984. In the updated study, the authors reported a statistically significant excess of lung cancer. In discussing this excess, the authors concluded that the interpretation of the information was difficult since there was no relationship between the excess of lung cancer and the length of time since first exposure to the fibrous glass manufacturing environment. Two recent case-control studies have addressed the lung cancer mortality of FG and slag wool production workers. Chiazze et al. (1992) have investigated the potential impact of confounding factors such as smoking and other occupational exposures for workers at the oldest and largest US fiber glass manufacturing facility. In particular, Chiazze helped clarify the heavy smoking patterns in those workers and verified the large impact that smoking has on their lung cancer experience. Wong et al., (1991) investigated the potential impact of smoking on the lung cancer deaths at nine US slag wool manufacturing plants. Wong also found heavy smoking among the slag wool workers and advanced the understanding of the modest increase in lung cancer seen in the historical cohort studies cited above. Users of MMVFs generally have experienced mixed exposures, making the study of any potential health effects of MMVF difficult, if possible at all. For example, in a study of Swedish construction workers, Engholm et al. (1987) discussed the difficulty caused by overlapping of reported exposures to asbestos and MMVFs. In addition, essential employment and exposure histories for users of MMVFs are lacking. The mortality studies of FG workers, while showing a small but statistically significant increase in lung cancer, have failed to show any consistent relationship with exposure to FG (i.e. no dose-response relationships have been found). It is recognized that uncontrolled occupational and/or non-occupational confounding factors may be associated with the slight increase. The IARC review (IARC, 1988) concluded that there is ‘inadequate evidence’ for carcinogenicity in humans. Other reviews have reached similar conclusions. In addition, reports subsequent to the IARC review have further clarified potential confounding factors and, if anything, shown weaker evidence of a relationship between exposure and lung cancer. The cohort mortality studies of rock wool and slag wool workers have shown a somewhat larger statistically significant excess of lung cancer deaths, but have also provided no clear dose-response relationship with

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fiber exposure. While the IARC review (IARC, 1988) concluded that there is ‘limited evidence’ for carcinogenicity in humans, reports subsequent to the IARC review have further clarified potential confounding factors and, if anything, shown weaker evidence of a relationship between exposure and lung cancer. Experimental studies Toxicologic studies of MMVFs have been conducted in both in vitro and in vivo systems. In addition, the physical and chemical characteristics thought to correlate with toxicity have been examined. The in vitro studies have been conducted using cells from the lungs of animals as well as bacterial and cell lines. Two categories of whole-animal studies have been reported: studies using artificial methods to implant high concentrations of fibers in the abdomen, pleura or trachea of animals; and inhalation studies of maximum tolerated doses and multiple dose levels of fibers. Cell culture studies The use of cell culture systems for studying the toxic effects of fibers has been recently reviewed (Hesterberg et al., 1993a). A number of studies have shown that fiber length and diameter are important in determining the toxicity of mineral fibers of various chemical compositions to cells grown in culture (Chamberlain et al., 1979, Tilkes and Beck, 1980; Hesterberg and Barrett, 1984; Hesterberg et al., 1993a; Hart et al., 1994). Chemical composition has also been shown to be critical to the toxicity of fibers to rat tracheal epithelial cells (Ririe et al., 1985) and human bronchial epithelial cells grown in culture (Kodama et al., 1993). MMVFs have also been shown to induce neoplastic transformation (Hesterberg and Barrett, 1984; Poole et al., 1986) and genetic damage to cells in culture (Sincock and Seabright, 1975; Oshimura et al., 1984). Cell culture models are important for understanding the mechanisms of fiber toxicity and, with further development, have potential for use as part of a battery of shortterm screening tests to assess the toxic and tumorigenic potential of mineral fibers. However, it was recently shown that cytotoxicity of different compositions MMVFs to Chinese hamster ovary (CHO) cells in culture did not correlate with the in vivo toxicity of theses MMVFs (Hart et al., 1994). This may be related to CHO cells being aneuploid, preneoplastic and not a normal target cell for fiber toxicity in vivo. Future in vitro studies of MMVF toxicity should focus on the use of cell types that represent the relevant target tissues, and cells should be as close to normal as possible.

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Implantation studies Using various types and dimensions of fibers, researchers have studied the effects of ‘artificially’ exposed animals by surgically implanting fibrous material in the pleural (chest) and abdominal cavities of laboratory animals, and by injecting fibers directly into the trachea (Pott et al., 1987; Stanton et al., 1981). Those studies have shown that high levels of most fibrous materials of certain dimensions, regardless of their physical or chemical makeup, can induce tumors in laboratory animals. From these study results, scientists have also hypothesized that biological activity correlates with fiber length and diameter, since ‘long, thin’ fibers are the most active. The actual chemical composition appears to play only a minor role, if any, in such ‘artificial exposure’ experiments (Stanton et al., 1981). Injection of fibers bypasses the normal defense mechanisms of the lung and can produce abnormal fiber distribution, fiber clumping, and overload doses (McClellan et al., 1992). Furthermore, when fibers are injected into the pleura or peritoneum of an animal, leaching, degradation, fragmentation or any other transformations are unlikely to be the same as after inhalation. The weaknesses of intracavitary injection studies of fibrous materials limit their relevance for human risk assessment (IPCS, 1988; Vu, 1988; Dement et al., 1990; WHO, 1992; McClellan et al., 1992). Recent animal inhalation studies of MMVFs Inhalation is the only natural route of exposure for fiber entry and distribution to the target organs in man. Animal inhalation studies are more relevant than intracavity administration studies for risk assessment because the exposure conditions of inhalation experiments more closely approach the circumstances of human exposure. Background In June 1988, a series of inhalation studies was initiated at Research and Consulting Company (RCC) in Geneva, Switzerland, to evaluate the biological effects of different compositions of MMVF. These included RCFs, common insulation fiber glass, and rock and slag wool fibers. These studies used recently perfected state-of-the-art technologies for fiber sizeseparation, fiber lofting and nose-only inhalation exposure. A more detailed description of the techniques used and the results from these studies are found elsewhere (Hesterberg et al., 1991, 1993b; Mast et al., 1995 McConnell et al., 1994). The animal models selected were those with demonstrated capacity to develop asbestosrelated disease following inhalation exposure. The studies were conducted in accordance with

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standard techniques for chronic toxicity/carcinogenicity studies, including dose and latency considerations. Rats were exposed for 2 years and hamsters for 18 months. The animals were observed for their lifetime or until 20% survival of the test group was reached. Positive and shamexposed negative controls were included in the protocol. Fiber aerosols In designing these animal inhalation studies, the techniques of fiber preparation, aerosolization, exposures, measurement, quantification and determination of actual target organ dose were critical factors. Fiber dimensions that permitted deposition into the distal lung regions (i.e. respirable fibers) for the model used were selected. The characteristics of the fiber aerosol in actual work areas for man was an important consideration in determining experimental exposure. For example, an average fiber size of 1×20 µm has been measured during simulated RCF work practices. The critical need to use fibers pre-selected for their size and to verify the actual size distributions of the fiber exposure aerosol was met throughout the study. Non-fibrous particles (shot) in the aerosol were reduced to the maximum extent possible. Furthermore, fiber preparation, handling and aerosolization did not alter the physical-chemical characteristics of the fiber, since as will be discussed later, these are known to be critical determinants of fiber toxicity. Nose-only rather than whole-body exposure was used for several reasons, including the impossibility of preparing the huge quantities of specially sized fibers that would be required for 2 years of whole-body exposure. Additionally, nose-only exposure levels permitted better control of exposure levels and host entry. Selection of exposure concentrations It was important that at least three exposure concentrations be used in the chronic inhalation study in order to assess the dose-response relationships of any induced changes. The highest concentration selected was the ‘Maximum tolerated dose’ (MTD), while lower concentrations were 50 per cent of the MTD and multiples of the projected occupational and environmental exposure levels. Experimental design, time lines Groups of three or six randomly selected animals from each exposure group were killed at 3, 6, 12, 18 and 24 (rats only) months to follow the progression of histopathological changes and to determine lung fiber burdens. An additional six ‘recovery’ animals were removed from each

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exposure group at 3, 6, 12 and 18 (rats only) months and held without further treatment until the end of the exposure period, when they were killed to assess progression or regression of lung lesions and lung retention and clearance of fibers after cessation of exposure. To assure quality control, the lofting technique and exposure level were consistently monitored during the study by both gravimetric measurement and fiber counting techniques. The terminal sacrifice was carried out when only 20 per cent of the animals survived. A complete necropsy was performed on each animal. Gross pathological examination and diagnoses were performed using a dissecting microscope. Uniform sections of the left lung and right diaphragmatic lobe were embedded in paraffin, cut at a thickness of 4 mm and replicate sections were routinely stained with hematoxylin and eosin (H&E) and Masson-Goldner’s trichrome stain for collagen staining to assess the presence of lung fibrosis. In addition, sections were made from all grossly visible lesions from that and other portions of the lung. Proliferative lesions of the pulmonary parenchyma were designated as bronchoalveolar hyperplasia, pulmonary adenoma or adenocarcinoma. Other types of lesions, including those in the pleura were noted where appropriate. All research and analyses were conducted using good laboratory practices. Lung fiber burden Immediately after necropsy, the infracardiac lobe of each animal’s lung was removed and frozen for later analysis of lung fiber burden. To recover fibers from the lung, the tissue was rapidly dehydrated with acetone and ashed using a low-temperature process. Recovered fibers were dispersed in distilled water and examined using scanning electron microscopy. Number, dimensions and other physical characteristics of the inhaled lung fibers were determined, and reported as fibers per mg of dry lung weight. Results from recent animal inhalation studies of MMVFs Refractory ceramic fibers In the first RCC studies, rodents were exposed to the MTD of the sizeselected RCF test fiber, 30 mg m−3 and approximately 200–250 fibers cm −3. Rats were exposed for 6 h per day, 5 days a week to aerosols containing one of four different types of RCF: kaolin, RCF 1; zirconia, RCF 2; high purity kaolin, RCF 3; and ‘after service’ (a kaolin based ceramic fiber containing 27% crystalline silica that had previously been exposed to high temperature), RCF 4. Hamsters were exposed to only kaolin RCF fibers. Positive controls (chrysotile asbestos) and negative controls (filtered air)

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Table 7.1 Summary of lung pathology findings in RCF hamster inhalation study

WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter <3 µm. Animals at risk are defined as all animals treated until the discovery of the first neoplasm of interest (i.e. lung tumors or pleural mesotheliomas) or for 1 year, whichever occurs first (the earliest neoplastic finding occurred at 9 months in this study). c Statistically significant increase over negative controls (p<0.001) using Fisher’s exact test. a

b

were included in both the rat and the hamster studies. Interim sacrifices in the RCF hamster study revealed lung fibrosis beginning at 9 months and development of malignant mesothelioma in 42 of 112 hamsters (38 per cent) exposed to RCF (Table 7.1). The rat MTD inhalation study revealed mesotheliomas in all groups of RCF exposed rats and significant increases in lung tumors in RCF1 and RCF3 compared to air breathing controls (Table 7.2). An additional rat chronic inhalation study was conducted, using multiple concentrations of RCF1. In this study one half the MTD (16 mg m−3, 120 f cm−3) as well as 9 mg m−3 (75 f cm−3) and 3 mg m−3 (26 f cm−3) were tested. The lung pathology resulting from exposure to different levels of RCF 1 is shown in Table 7.3. Data from the RCF MTD study in rats for the 30 mg m−3 dose of RCF and from the 10 mg m−3 of chrysotile asbestos positive control are also included in the table for comparison. No significant increase in lung tumors was observed in rats exposed to RCF 1 doses below 30 mg m−3 (187 f cm−3), although lung fibrosis and a single mesothelioma were observed at 9 mg m−3 (75 f cm−3) (Table 7.3). Fiber glass Groups of rats were exposed as in the RCF studies, to size-selected Schuller 901 fiber glass (MMVF 10) or Certain Teed B fiber glass (MMVF 11). The MMVFs used in these studies were size-selected from commercial insulation wool to have an average diameter of approximately 1 µm and an average target length of 20 µm. They were therefore, as with the RCF test fibers, analogous to the dimensions of fibers found in workplace air and also rat respirable. A summary of the lung tumor findings is shown in Table 7.4. Lung tumors were found in the negative control group as well as

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Table 7.2 Summary of lung pathology findings in RCF MTD (30 mg m−3) rat inhalation

WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter <3 µm. Animals at risk are defined as all animals treated until the discovery of the first neoplasm of interest (i.e. lung tumors or pleural mesotheliomas) or for 1 year, whichever occurs first (the earliest neoplastic finding occurred at 1 year in this study). c Statistically significant increase over negative controls (p<0.001) using Fisher’s exact test. a

b

in most of the treated groups. There are no statistically significant differences in any of the exposure groups when compared to the air controls. Dose response tests on the final tumor incidences showed no significance. Heterogeneity among the groups is in evidence, and the overall interpretation of the results indicates no relationship between exposure to glass fibers used in this study and disease of any type. The background air control lung tumor level found in this study (3.1%) is similar to the average background control level of 2.9% reported in a previous review of Fischer 344 Rat chronic studies (Solleveld et al., 1984). Background lung tumor levels that ranged from 0 to 8 per cent for bronchoalveolar adenomas and from 0 to 6 per cent for bronchoalveolar carcinomas in this strain of rats has been reported (Boorman and Eustis, 1990). This indicates that the tumor levels within the fibrous glass exposure groups in the present study agree closely with this background control range. Slag wool and rock wool Rats were exposed as in the previous studies to 3, 16 or 30 mg m−3 (30, 131, 213 f cm−3, respectively) of size separated slag wool fiber or to similar concentrations (34, 150 and 243 f cm−3 respectively) of rock wool fiber. Negative controls (filtered air) and positive controls exposed to crocidolite

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Table 7.3 Combined summary of lung pathology findings: chrysotile and RCF1 from RCF MTD study in rats; and RCF multidose study in rats

WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter <3 µm. Animals at risk are defined as all animals treated until the discovery of the first neoplasm of interest (i.e. lung tumors or pleural mesotheliomas) or for 1 year, whichever occurs first (the earliest neoplastic finding occurred later than 1 year in this study). c From multidose study. d From MTD study—statistical comparison with concurrent controls (see Table 7.2). e No dose response trend (negative controls included) was observed testing for trend on just the RCF Multidose study using an exact algorithm for the CochranArmitage test. A significant dose response trend (p<0.001 using combined negative controls) was observed testing for trend on all RCF1 treatment groups using an exact algorithm for the Cochran-Armitage test. f Statistically significant increase over negative controls (p<0.001) using Fisher’s exact test. a

b

asbestos (10 mg m−3) were included in the study. The crocidolite exposure had to be stopped at 10 months due to excessive mortality resulting from lung toxicity. The results are summarized in Table 7.5. Crocidolite exposure resulted in lung fibrosis, a significant increase in lung tumors, and a single mesothelioma. Rock wool, but not slag wool, exposure at 16 and 30 mg m−3 resulted in minimal lung fibrosis. However, neither rock wool nor slag wool exposure resulted in mesotheliomas or a significant increase in lung tumors. Lung burden analyses To more fully understand the critical fiber characteristics responsible for the differences in MMVF toxicity, the characteristics of the fibers in the lungs of the exposed animals were compared. Since RCF was the only MMVF type that induced lung tumors and this effect only occurred at the

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Table 7.4 Summary of lung pathology findings in fibrous glass inhalation study in rats

WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter <3 µm. Animals at risk are defined as all animals treated until the discovery of the first neoplasm of interest (i.e. lung tumors or pleural mesotheliomas) or for 1 year, whichever occurs first (the earliest neoplastic finding occurred later than 1 year in this study). c No dose response trend (negative controls included) was observed when tests for trend were made using an exact algorithm for the Cochran-Armitage test. d From RCF MTD study—statistical comparison with concurrent controls (see Table 7.2). e Statistically significant increase over negative controls (p< 0.001) using Fisher’s exact test. a

b

highest dose (30 mg m−3), fiber characteristics only at this highest dose were compared. The length and diameter distributions of the different MMVF types recovered from rats exposed for 13 weeks are shown in Figure 7.1. Figure 7.2 illustrates the numbers of lung fibers in each of three length-categories recovered at various time points during the exposure phases of the MMVF studies, The WHO (World Health Organization) criteria define a respirable fiber as having an aspect ratio of ≥3:1, a length >5 µm, and a diameter <3 µm (WHO, 1985). The numbers of WHO fibers and fibers >10 µm in length in the lung were similar for each of the different MMVF types (Figures 7.2(a) and 7.2(b)). However, greater numbers of long fibers (>20 µm. long) were found in the lungs of rats exposed to RCF 1 and MMVF 21 (rock wool) than for the other fiber types (Figure 7.2(c)). Even though lung levels of long MMVF 21 fibers were higher than long RCF 1 fibers, lung fibrosis occurred much later for MMVF 21 (18 vs 6 months for RCF 1) and no mesotheliomas or significant increase in lung tumors were observed for MMVF 21. This indicates that the lung pathogenic potential of a fiber may be determined by more than dose and dimension.

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Table 7.5 Summary of lung pathology findings in rock and slag wool inhalation study in rats

WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter <3 µm. Animals at risk are defined as all animals treated until the discovery of the first neoplasm of interest (i.e. lung tumors or pleural mesotheliomas) or for 1 year, whichever occurs first (the earliest neoplastic finding occurred later than 1 year in this study). c Exposure to crocidolite asbestos had to be discontinued after 10 months due to excessive mortality. d No dose response trend (negative controls included) was observed when tests for trend were made using an exact algorithm for the Cochran-Armitage test. e Statistically significant increase over negative controls (p=0.001) using Fisher’s exact test. a

b

In a previous paper on this series of studies, it was reported that fiber glass and RCF reached similar levels in the lung during continuous exposure and appeared to clear from the lung at comparable rates (Hesterberg et al. 1994). Fibers recovered from the lungs of animals exposed for only 13 weeks and held without further exposure for 91 weeks (recovery animals) were analyzed for chemical change using energy dispersive spectroscopy (EDS) with Scanning Electron Microscopy. These analyses of recovery lungs showed that much of the alkalis and alkaline earth components had leached from the fiber glass over time. However, only a slight change in RCF chemistry was observed. These findings indicate that the leaching of fibers and the resultant change in chemical composition, especially the surface chemistry, may be an important determinant of the biological activity of MMVFs. The importance of chemical composition to the toxic potential of fibers has been recently reviewed (Guthrie and Mossman, 1993). In addition, fibers that are more rapidly leached of their alkalis and alkaline earths may be more readily broken in the lung. This might explain why fibers with more leachable compositions show the most rapid

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disappearance of long fibers. More studies are required to determine if a fiber’s ability to be leached is a critical determinant of its ultimate toxicity to the lung.

Figure 7.1 Length distributions (a) and diameter distributions (b) of fibers from the lungs of rats exposed for 13 weeks to the five different MMVFs in the chronic inhalation studies. To permit clearance of the upper airways, rats were killed a

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minimum of 24 h after the exposure was stopped; the right accessory lobe was frozen and later low temperature ashed for fiber recovery. Fiber lengths were determined using phase contrast optical microscopy, while fiber diameters were determined using scanning electron microscopy.

Results from previous MMVF inhalation studies The results from two previous RCF inhalation studies (Davis et al., 1984; Smith et al., 1987) differ from the more recent RCC studies presented here. Davis et al., (1984) reported RCF exposure of rats resulted in an average of 5 per cent pulmonary fibrosis, pulmonary tumors in eight of 48 rats, and one peritoneal mesothelioma. The lower fibrosis and tumor response in the Davis study may have resulted from the lower exposure concentration used; 8.4 mg m−3 compared to 30 mg m−3 in the present study. In addition, the use of fibers that were not presized, the use of whole-body exposure, or the fiber generation technique, which may have crushed some of the fibers, may account for the lack of consistency with the present study. Smith et al., (1987) exposed hamsters and rats to RCF at 200 f cm−3, 6h a day, 5 days a week, for 24 months. The rat study showed no significant increase in neoplasms and minimal pulmonary fibrosis in 22% of the exposed animals. In the hamster study, RCF produced only one mesothelioma in 50 animals and no fibrosis was observed. It is difficult to explain why there was little response to RCF in the Smith studies, but it may be related to the different aerosol and exposure technology used or to the low exposure level; 12 mg m−3 compared to 30 mg m−3 in the present study. Previous inhalation studies of FG using rodents agree with the findings of the RCC studies. Fiber glass has been tested by inhalation in guinea pigs (Gross et al., 1970), hamsters (Lee et al., 1981; Smith et al., 1987), and rats (Gross et al., 1970; Lee et al., 1981; McConnell et al., 1984; Wagner et al., 1984; Mitchell et al. 1986; LeBouffant et al., 1987; Muhle et al., 1987; Smith et al., 1987). None of these studies identified a significant increase in either fibrosis or neoplasms following glass fiber inhalation in spite of FG lung burdens in excess of several hundred thousand fibers per mg dry lung tissue. In three of the above studies, the chronic inhalation toxicity of rock and slag wool were also examined (Wagner et al., 1984; LeBouffant et al., 1987; Smith et al., 1987). As was seen with fibrous glass, all three studies demonstrated no tumorigenic response by this route of exposure.

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Figure 7.2 (continued over) Lung burdens of (a) WHO fibers, (b) fibers >10 µm in length, and (c) fibers >20 µm in length per mg dry lung tissue from the lungs of rats continuously exposed to 30 mg m−3 of the five different MMVFs. Rats were killed at least 24 h after the exposure was stopped, the right accessory lobe was frozen and later low temperature ashed for fiber recovery.

Industrial hygiene studies RCF Industrial hygiene monitoring data obtained on a regular basis at locations

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Figure 7.2 Continued where RCF products are manufactured show that exposures are generally below 1.0 f cm−3, typically below 0.2 f cm−3. In a recent study, RCF levels during various end-user operations ranged from 0.12 to 1.55 f cm−3 with an overall mean and SD of 0.74±0.49 f cm−3 (Lees et al., 1993b). Other end-user studies have indicated that RCF exposures can exceed 5 f cm−3 or higher if appropriate engineering controls and work practices are not followed (Schuller, 1985–1988). Fiber glass Recently, studies which examined human aerosol exposure to fiber glass in manufacturing, installation and removal, and in ambient air were reviewed (Hesterberg and Hart, 1994). In most cases, human exposures to airborne fiber glass during manufacturing and installation fell well below the OSHAproposed permissible exposure limit (PEL) of 1 f cm−3 air (OSHA, 1992). Airborne fiber concentrations during FG manufacturing operations are typically less than 0.2 f cm−3, with the majority being less than 0.1 f cm−3. Exceptions include manufacture of finer diameter fiber glass and blowing installation of loose fiber glass that is either milled or lacks binder. Airborne levels averaging greater than 1 f cm−3 have been reported in the production of finer diameter fiber glass (TIMA, 1990), while blowing installation of loose fiber glass without binder resulted in a task length average (TLA) of 7.67 f cm−3, and an 8-h TWA of 1.96 f cm−3 (Lees et al., 1993a). Blowing installation of loose mineral wool also resulted in higher aerosol levels; a

108 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS

TLA of 1.94 f cm−3 and an 8-h TWA of 0.97 f cm−3 (Lees et al., 1993a). Removal of fiber glass insulation created an aerosol of 0.042 f cm−3 (Jacob et al., 1993). Fiber concentrations of 0.004 f cm−3 were reported for buildings recently insulated with FG (Jacob et al., 1992). However this figure includes all types of fibers as it was obtained using optical microscopy. The background level prior to fiber glass installation was 0.001 f cm−3. Ambient environmental exposures to airborne vitreous fibers were extremely low; exposure levels of product-related vitreous fibers reported for outdoor air was 0.0007 f cm−3 (Tiesler and Draeger, 1994). In addition to manufacturing and field use surveys, release of fibrous glass during actual use of products, particularly fiber released from air filter media, has been monitored. To determine possible exposure of building occupants to fibrous glass, ambient air was sampled in a number of public buildings in which fibrous glass air filtration products had been installed. These evaluations showed no significant release of fibers from the filters (Balzer et al., 1971; Cholak and Schafer, 1971). To evaluate the efficiency of fibrous glass filter blankets, several high volume air samples were collected at various points in the ductwork of a large office complex at the intake and the exhaust prior to changing the filter media, and at the exhaust 23 days after installation of the new filter. Analyses of the samples using electron microscopy indicate little initial fiber release which decreases rapidly thereafter to the limit of detection (Schuller, 1987). Rock and slag wool Airborne concentrations of dust and fibers reported from US mineral wool plants is generally higher than in US glass wool facilities. This includes both airborne fibers and total particulate matter. Fiber levels reported ranged from 0.01 to 1.4 f cm−3, compared with 0.1–0.3 f cm−3 for glass wool. Total particulate matter sample results ranged from 0.05 to 23.6 mg m−3 in the mineral wool facilities and 0.09–8.48 mg m−3 for glass wool (Esmen et al., 1980). Comparison of Human MMVF exposures used in the recent rat chronic inhalation studies When using animal inhalation studies for assessment of potential risk to human health of airborne fibers, it is critical to demonstrate that the characteristics and concentrations of the experimental fiber aerosols are comparable with those in human exposure situations. It is also important for risk assessment that the actual target organ dose in the animal model reach or exceed that found in exposed humans. To illustrate, consider levels of fiber glass published in a number of recent reports. A qualitative

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Table 7.6 Representative airborne levels of fiber glass in workplace and rat inhalation study

a

Outdoor data from Tiesler and Draeger (1994). Product-related fibers counted using NIOSH A Rules. b Data from Jacob et al., (1993). Airborne levels resulting from manufacturing operations using FG insulation. c Data from Lees et al., (1993a). Installation of residential insulation. d Jacob et al. (1992) reported that levels returned to background within hours after Batt installation. All other data are averages from the various studies herein cited.

and quantitative comparison was made of the aerosol and lung fibers in the rat inhalation study with those in various human exposure situations (Hesterberg and Hart, 1994). A comparison of the reported aerosol fiber levels in various human settings with those used in the rat inhalation study is shown in Table 7.6. FG levels in the rat aerosol were more than five orders of magnitude higher than the reported level for outdoor air, and at least three orders of magnitude higher than for average airborne levels for many occupational settings (e.g. over 2000fold higher than FG batt installation). The rat aerosol was 75-fold more concentrated than the highest reported average TWA for airborne fiber levels in an occupational setting, i.e. blowing installation of unbound fiber glass (the potential for higher airborne levels has been recognized for some time, and recommended work practices call for the use of respirators in such circumstances). Despite the range in products and occupational settings, fiber dimensions in most of the human exposures examined were fairly similar to those found in the rat inhalation study aerosol (Hesterberg and Hart, 1994). The fiber dimensions of aerosolized rock and slag wool collected from workplace air during the installation of batts or blowing of loose fibers have similar mean diameters to that of fiber glass (1.0–1.6 µm). However, the mean lengths appear to be greater (30–50 µm) than for most workplace samples of fiber glass. Hesterberg and Hart (1994) also compared the lung burdens of rats exposed in the recent fiber glass inhalation study in rats with lung burdens found in workers involved in MMMF (primarily FG) manufacturing (McDonald et al. 1990). As shown in Table 7.7, rat fiber glass lung burdens vastly exceeded that of the workers reported by McDonald et al.,

110 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS

which was not significantly elevated above reference levels. Fibers per mg dry lung for the rat after lifetime exposure was >4000-fold higher than for the fiber glass worker, average exposure 11 years (the average time from last employment in MMMF manufacturing and death was 12 years). Lung fiber dimensions in the rat study were comparable to those of fibers recovered from the lung tissue of fiber glass manufacturing workers. From these comparisons, it can be concluded that the exposure levels used in the recent rat inhalation studies unequivocally achieved the goal of the studies to exceed human exposures by several orders of magnitude. Summary and conclusions MMVFs are among the most studied commercial products due to their widespread use and the concern for potential health effects of respirable fibers. In recent animal inhalation studies RCF produced lung fibrosis, mesotheliomas, and significant increases in lung tumors. However, it is believed that any potential cancer risk from RCF exposure can be minimized, if not eliminated, because of the small number of workers exposed and the ability to use respiratory protection and engineering controls to limit worker exposure. Both human and animal inhalation studies have shown no association between fiber glass exposure and disease. Although high exposure levels of rock wool (several orders of magnitude higher than most reported workplace exposures) produced minimal lung fibrosis in rats, no mesotheliomas and no significant increase in lung tumors were observed. Slag wool produced no fibrosis or increase in tumors in the animal studies. The cohort mortality studies of rock wool and slag wool workers have also provided no clear dose-response relationship with fiber exposure. Results from the combined animal inhalation studies showed that differences in lung fiber burdens and lung clearance rates could not explain the differences observed in the toxicologic effects of MMVFs. These findings clearly indicate that dose, dimension and durability (i.e. the persistence of fibers in the rat lung) are not the only determinants of fiber toxicity; chemical composition and the surface physicochemical properties of the fibers may also play an important role. Exposure levels from animal inhalation studies were at least three orders of magnitude higher than for average airborne levels reported for many occupational settings.

b

Lung fibers: for humans, NIOSH A rules; for rats, total fibers (all fibers length/diameter >3:1). For humans, NIOSH A rules; for rats, WHO respirable fibers, comparable to A rules because there were no diameters >3 µm in rats. c Hesterberg et al., (1993b), Rat fiber exposure was 5 days week•1, 6 h day•1 for lifetime (2 years). d McDonald et al., (1990). Negative controls had not worked with FG and were matched with each FG worker for age and locale. e Occupational exposures averaged 11 years, followed by average of 12 years without exposure prior to death. f 101 were FG workers; 11 were mineral wool workers. g Not reported.

a

Table 7.7 Reported lung fiber levels from fiber glass workers and rat inhalation study

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112 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS

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HESTERBERG, T.W., Vu, V., CHASE, G.R., MCCONNELL, E.E., BUNN, W.B. and ANDERSON, R., 1991, Use of animal models to study man-made fiber carcinogenesis, in Brinkly, B., Lechmer, J. and Harris, C. (Eds) Cellular and Molecular Aspects of Fiber Carcinogenesis, pp. 183–205, Cold Spring Harbor New York: Cold Spring Harbor Laboratory. HESTERBERG, T.W., HART, G.A. and BUNN, W.B., 1993a, In vitro toxicology of fibers: mechanistic studies and possible use for screening assays, in Warheit, D. (Ed) Fiber Toxicology New York: Academic Press. HESTERBERG, T.W., MIILLER, W.C., MCCONNELL, E.E., CHEVALIER, J., HADLEY, J., BERNSTEIN, D.M., THEVENAZ, P. and ANDERSON, R., 1993b, Chronic inhalation toxicity of size-separated glass fibers in Fischer 344 rats, Fundam. Appl. Toxicol, 20, 464–76. HESTERBERG, T.W., MCCONNELL, E.E., MIILLER, W.C., BERNSTEIN, D.M., MAST, R. and ANDERSON, R., 1994, Relationship between lung biopersistence and biological effects of man-made vitreous fibres after chronic inhalation in rodents, in Bignon, J., Saracci, R. and Touray, J.-C. (Eds) Biopersistence of Respir-able Synthetic Fibres and Minerals (in press). IARC, 1988, Monographs on the Evaluation of Carcinogenic Risks to Humans, Vol. 43, World Health Organization International Agency for Research on Cancer, Man-Made Mineral Fibres, Environmental Health Criteria 77, Geneva. IPCS, 1988, International Programme on Chemical Safety, Man-made Mineral Fibers, Environmental Health Criteria Document No. 77, Geneva: World Health Organization. JACOB, T.R., HADLEY, J.G., BENDER, J.R. and EASTES, W., 1992, Airborne glass fiber concentrations during installation of residential insulation, Am. Ind., Hyg. Assoc. J., 53, 519–23. JACOB, T.R., HADLEY, J.G., BENDER, J.R. and EASTES, W., 1993, Airborne glass fiber concentrations during manufacturing operations involving glass wool insulation, Am. Ind., Hyg. Assoc. J., 54, 320–6. KODAMA, Y., MANESS, S.C., IGLEHART, J.D., BOREIKO, C.J. and HESTERBERG, T.W., 1993, Cytotoxic and cytogenetic effects of asbestos to human bronchial epithelial cells in culture, Carcinogenesis, 14, 691–7. LEBOUFFANT, L., DANIEL, H., HENIN, J.P., MARTIN, J.C., NORMAND, C., THICHOUX, G. and TROLARD, F., 1987, Experimental study on long-term effects of inhaled MMMF on the lung of rats, Ann. Occup. Hyg., 31, 765–90. LEE, K.P., BARRAS, C.E., GRIFFITH, F.D., WARITZ, R.S. & LAPIN, C.A., 1981, Comparative pulmonary responses to inhaled inorganic fibers with asbestos and fiberglass, Environm. Res., 24, 167–91. LEES, P., BREYSSE, P., MCARTHUR, B., MILLER, M., ROONEY, B., ROBBINS, C. and CORN, M., 1993a, End user exposures to man-made vitreous fibers: I. Installation of residential insulation products. Appl. Occup. Environ. Hyg. 8(12), 1022–30. LEES, P., BREYSSE, P., MCARTHUR, B., MILLER, M., ROONEY, B., ROBBINS, C. and CORN, M., 1993b, End user exposures to refractory ceramic fiber insulation products, unpublished report. MARSH, G.M., ENTERLINE, P.E., STONE, R.A. and HENDERSON, V.L., 1990, Mortality among a cohort of U.S. man-made mineral fiber workers: 1985 followup, J. Occup. Med., 32, 594–604.

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MAST, R.W., MCCONNELL, E.E., ANDERSON, R., CHEVALIER, J., KOTIN, P., BERNSTEIN, D.M., THEVENAZ, P. GLASS, L.R. MILLER, W.C. and HESTERBERG, T.W., 1995, Studies on the chronic toxicity (inhalation) of refractory ceramic fiber in male Fischer 344 rats, Inhalat. Toxicol., 7, 425–67. McCLELLAN, R.O., MILLER, F.J., HESTERBERG, T.W., WARHEIT, D.B., BUNN, W.B. et al., 1992, Approaches to evaluating the toxicity and carcinogenicity of man-made fibers, Regulat. Toxicol. Pharmacol., 16, 321– 64. MCCONNELL, E.E., WAGNER, J.C., SKIDMORE, J.W. and MOORE, J.A., 1984, A comparative study of the fibrogenic and carcinogenic effects of UICC Canadian chrysotile B asbestos and glass microfibre (JM 100), in Biological Effects of Man-made Mineral Fibres (Proceedings of a WHO/IARC Conference), Vol. 2. pp. 234–52, Copenhagen: World Health Organization. MCCONNELL, E.E., KAMSTRUP, O., MUSSELMAN, R., HESTERBERG, T.W., CHEVALIER, J., MIILLER, W.C. and THEVENAZ, P., 1994, Chronic inhalation study of size-separated rock and slag wool insulation fibers in Fischer 344/N rats, Inhalat. Toxicol., 6, 571–614. MCDONALD, J.C., CASE, B.W., ENTERLINE, P.E., HENDERSON, V., MCDONALD, A.D., PLOURDE, M. and SEBASTIEN, P., 1990, Lung dust analysis in the assessment of past exposure of man-made mineral fibre workers, Ann. Occup. Hyg., 34(5), 427–41. MITCHELL, R.I., DONOFRIO, D.J. and MOORMAN, W.J., 1986, Chronic inhala-tion toxicity of fibrous glass in rats and monkeys, J. Am. Coll. Tox., 5, 545–75. MUHLE, H., POTT, F., BELLMANN, B., TAKENAKA, S. and ZIEM, U., 1987, Inhalation and injection experiments in rats to test the carcinogenicity of MMMF, Am. Occup. Hyg., 31, 755–64. NIOSH Manual of Analytical Methods, 3rd. ed, US Dept of Health & Human Services, Public Health Service, Centers for Disease Control. OSHA (Occupational Safety and Health Administration), 1992, Proposed Rules for Fibrous Glass, Including Refractory Ceramic Fibers, Federal Register, Vol. 57, No. 114, p. 26195. OSHIMURA, M., HESTERBERG, T.W., TSUTSUI, T. and BARRETT, C.J., 1984, Correlation of asbestos-induced cytogenetic effects with cell transformation of Syrian hamster embryo cells in culture, Cancer Res., 44, 5017–22. POOLE, A., BROWN, R.C. and ROOD, A.P., 1986, The in vitro activities of a highly carcinogenic mineral fibre—potassium octatitanate, Br. J. Exp. Pathol, 67, 289–96. POTT, F., ZIEM, U., REIFFER, F.J., HUTH, F., ERNST, H. and MOHR, U., 1987, Carcinogenicity studies of fibres, metal compounds and some other dusts in rats., Exp. Pathol, 32, 129–52. RIRIE, D.G., HESTERBERG, T.W., BARRETT, J.C. and NETTESHEIM, P., 1985, Toxicity of asbestos and glass fibers for rat tracheal epithelial cells in culture, in Beck, E.G. and Bignon, J. (Eds), In vitro Effects of Mineral Dusts, NATO ASI Series, Vol. G3, pp. 177–184, Berlin (West): Springer. SARACCI, R., SIMONATO, L., ACHESON, E.D., ANDERSEN, A., BERTAZZI, P. A., CLAUDE, J., CHARNAY, N., ESTEVE, J., FRENTZEL-BEYME, R.R., GARDNER, M.J., JENSEN, O.M., MAASING, R., OLSEN, J.H., TEPPO, L.,

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WESTERHOLM, P. and ZOCCHETTI, C., 1984, Mortality and incidence of cancer of workers in the man-made vitreous fibres producing industry: an international investigation at 13 European plants, Brit. J. Indust. Med. 41, 425–36. Schuller International, Inc., 1985–1988, Unpublished data from industrial hygiene surveys. Schuller International, Inc., 1987, Unpublished data from industrial hygiene surveys. SHANNON, H.S., HAYES, M., JULIAN, J.A. and MUIR, D.C.F., 1987, Mortality experience of Ontario glass fibre workers—extended follow-up, Ann. Occup. Hyg., 31, 657–62. SIMONATO, L., FLETCHER, A.C., CHERRIE, J., ANDERSEN, A., VERTAZI, P., CHARNAY, N., CLAUDE, J., DODGSON, J., ESTEVE, J., FRENTZELBEYME, R., GARDNER, M.J., JENSEN, O., OLSEN, J., TEPPO, L., WINKELMANN, R., WESTERHOLM, P., WINTER, P.D., ZOCCHETTI, C. and SARACCI, R., 1987, The International Agency for Research on Cancer historical cohort study of MMMF production workers in seven European countries, extension of the followup, Ann. Occup. Hyg., 31, 603–23. SINCOCK, A. and SEABRIGHT, M., 1975, Induction of chromosome changes in Chinese hamster cells by exposure to asbestos fibres, Nature, 257, 56–8. SMITH, D.M., ORTIZ, L.W., ARCHULETA, R.F. and JOHNSON, N.F., 1987, Long-term health effects in hamsters and rats exposed chronically to manmade vitreous fibers, Ann. Occup. Hyg., 31, 731–54. SOLLEVELD, H.A., HASEMAN, J.K. and McCoNNELL, E.E., 1984, Natural history of body weight gain, survival and neoplasia in the F344 rat, J. Nat. Cancer Inst., 72(4), 929–40. STANTON, J.F., LAYARD, M., TEGERIS, A., MILLER, E., MAY, M., MORGAN, E. and SMITH, A. 1981, Relation of particle dimension to carcinogenicity in amphibole and other fibrous minerals, J. Nat. Cancer Inst., 67, 965–75. TIESLER, H. and DRAEGER, U., 1994, Measurement and identification of insulation product-related fibres in constrast to ubiquitous fibres, in Proceedings from the Symposium on Indoor Air, Helsinki, January, 1993, in press. TILKES, F. and BECK, E.G., 1980, Comparison of length-dependent cytotoxicity of inhalable asbestos and man-made mineral fibres, in Wagner, J.C., (Ed.) Biological Effects of Mineral Fibres (IARC Scientific Publications No. 30), pp. 475–83, Lyon: International Agency for Research on Cancer. TIMA (Thermal Insulation Manufacturer’s Association), 1990, Health and Safety Aspects of Man-made Vitreous Fibres: Information, Data, Comments and Recommendations Regarding Occupational Exposure to Man-made Vitreous Fibers, Part 3 Glass Fibers, Sections III, IV, submitted 7/10/90 to the National Institute for Occupational Safety and Health in response to 55 Fed. Reg. 5073. Vu, V. (US Environmental Protection Agency, EPA), 1988, Health Hazard Assessment of Nonasbestos Fibers, Final Draft, Health and Environmental Review Division, Office of Toxic Substances. WAGNER, J.C., BERRY, G.B., HILL, R.J., MUNDAY, D.E. and SKIDMORE, J. W., 1984, Animal experiments with MMM(V)F—effects of inhalation and intrapleural inoculation in rats, in Biological Effects of Man-made Mineral

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8 Pulmonary Toxicity Studies with Man-made Organic Fibres: Preparation and Comparisons of Size-separated Para-aramid with Chrysotile Asbestos Fibres DAVID B.WARHEIT,1 MARK A.HARTSKY,1 CHARLES J.BUTTERICK2 and STEVEN R.FRAME1 1

DuPont Haskell Laboratory, Newark, DE, 2 Texas Tech Health Sciences Center Lubbock, TX Introduction

This study was designed to compare the pulmonary toxic effects of inhaled, size-separated preparations of chrysotile asbestos fibres with paraaramid fibrils at similar aerosol fibre concentrations. Chrysotile samples are known to have an abundance of short fibres, with mean lengths generally in the range of 2 µm. This is important to note because one of the critical factors influencing the pathogenesis of fibre-related lung disease is fibre dimension (Davis et al., 1986). As a consequence, attempts were made to selectively enhance the mean lengths of chrysotile asbestos fibres used in this inhalation toxicity study, in order to make reasonable comparisons between the two fibre-types. Methods General experimental design Groups of male Crl: CDBR rats (7–8 weeks old, Charles River Breeding Laboratories, Kingston, New York) were used to assess the pulmonary effects of 2-week inhalation exposures to size-separated preparations of Kevlar® para-aramid fibrils or chrysotile asbestos fibres. Animals were exposed 6 hr day−1, 5 days week−1 for 2 weeks. For this study, Kevlar® was utilized as a representative para-aramid fibril. The two commercial types of para-aramid fibres are Twaron®, made by Akzo, and Kevlar®, made by DuPont. Following exposure, the lungs of p-aramid or chrysotile-exposed animals and age-matched sham controls were subsequently evaluated by bronchoalveolar lavage fluid analysis at 0 h, 5 days, 1 and 3 months postexposure. The lungs of additional animals were evaluated for biodurability, pulmonary clearance, pulmonary histopathologic lesions and

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lung and mesothelial cell proliferation at 0 hrs, 5 days 1, 3, 6 and 12 months postexposure. Fibre preparation and inhalation exposure Ultrafine Kevlar® p-aramid fibrils were supplied by DuPont Fibres. A special preparation of respirable p-aramid fibrils which had been prepared for the 2-year inhalation study (Lee et al., 1988) was utilized for this study. Bulk Canadian chrysotile asbestos fibres were obtained from Mr John Addison of the Institute of Occupational Medicine in Edinburgh, Scotland. Attempts were made to size-separate the bulk fibre preparation (i.e. selectively enhance the percentages of long fibres while removing the short fibres) by placing the fibres in a rotating sieve shaker and sieving through a series of metal mesh screens. The fraction containing the longer fibres (and a number of short fibres) was collected and generated for inhalation studies; fibres were collected on a filter and dimensional analysis (i.e. length and diameter assessments) was performed using scanning electron microscopy. The results showed that this technique was partially successful as the median and mean lengths of fibres were increased from 3 and 5 µm, respectively, in the original bulk sample to 6 and 9 µm in the generated sample preparation. The median lengths and diameters of p-aramid fibrils used in the study were 9 µm and 0.3 µm, respectively. The methods for aerosol generation of p-aramid fibrils have previously been reported (Warheit et al., 1992). Final mean fibre concentrations for the p-aramid exposures were 772 and 419 f cm−3. Aerosols of chrysotile asbestos fibres were generated in a similar manner, i.e. with a binfeeder and baffles, but without the microjet apparatus. Final mean fibre concentrations for the chrysotile asbestos exposures were 782 and 458 f cm−3. Fibre lung burdens were quantified from digested lung tissue of animals sacrificed immediately after the end of the 2-week exposure. Pulmonary lavage and biochemical assays on lavaged fluids Bronchoalveolar lavage procedures, cell quantification, and biochemical assays were conducted according to methods previously described (Warheit et al., 1984a, 1992). In addition, the methods for measuring lactate dehydrogenase (LDH), N-acetyl-β-glucosaminidase (NAG), and alkaline phosphatase (ALP) and protein in BAL fluids have been reported (Warheit et al., 1992).

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Lung dissection, tissue preparation and pulmonary cell proliferation The lungs of rats exposed to p-aramid and chrysotile asbestos fibres for 2 weeks were prepared for light microscopy by airway infusion using methods previously reported (Warheit et al., 1984b, 1991). Pulmonary cell proliferation experiments were designed to measure the effects of fibre inhalation exposure on terminal bronchiolar, proximal lung parenchymal (i.e. alveolar duct bifurcations and adjacent areas), subpleural and visceral pleural, and mesothelial cell turnover in rats following 2-week exposures. Groups of sham and fibre-exposed rats were given a 2-h pulse immediately after exposures, as well as 5 days, 1, 3, 6 and 12 months (still in progress) postexposure with an intraperitoneal injection of 5-bromo-2′deoxy-uridine (BrdU) dissolved in a 0.5N sodium bicarbonate buffer solution at a dose of 100 mg kg−1 body weight as previously described (Warheit et al., 1992). In addition, sections of duodenum served as a positive control. For each treatment group, there were immunostained nuclei in airways (i.e. terminal bronchiolar epithelial cells), lung parenchyma (i.e. epithelial, interstitial cells or macrophages), subpleura and visceral pleura, and mesothelial cells. All regions were counted by light microscopy at ×1000 magnification. Statistics were carried out using a two-tailed Students t test on a Microsoft Excel software program. Fibre recovery from lung tissue Para-aramid fibrils were recovered from the lungs of exposed rats using a diluted 1.3% hypochlorite (Clorox bleach) solution. The results of validation studies in our laboratory demonstrated that the dilute Clorox solution (10 min digestion) was more effective in digesting lung tissue than the KOH method that we had previously reported (Warheit et al., 1992). Chrysotile asbestos fibres were recovered from the lungs of exposed rats by incubating the lung tissue with a 5.25% hypochlorite solution for 3 h. Subsequently, the filters containing fibres recovered from lung tissue were mounted and prepared for phase-contrast light microscopy (for counting) and for scanning electron microscopy (for fibre dimensional analysis), according to methods previously described (Warheit et al., 1992). Results Size-separation methods for chrysotile asbestos fibres The results from size-separation attempts showed that there was a shift in the distribution of fibre lengths from shorter fibres to longer fibres (Figures 8.1(A)– (C)). Count median lengths of chrysotile asbestos fibres

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were increased from 3 µm in the original generated sample to 6 µm in the size-separated sample. In comparison to the chrysotile asbestos sample, there was a significantly greater proportion of long p-aramid fibrils which were used in the inhalation study with median lengths >9 µm. Lung burden analysis Although the aerosol fibre concentrations were similar throughout the study (p-aramid high conc.=772 f cm−3, chrysotile high conc.=782 f cm−3; p-aramid low conc.=419 f cm−3, chrysotile low conc.=458 f cm−3), measurement of lung fibre burdens from digested lung tissue at time 0 (i.e. immediately after exposure) demonstrated a substantial difference in lung burden between the two fibre-types as measured by phase contrast optical microscopy (PCOM). The mean lung fibre (>5 µm) burden from 3 rats/ dose group exposed to chrysotile asbestos was 3.7×107 (±7.4×106) fibres/ lung for the high dose group and 1.3×107 (±4×106) fibres/lung for the low dose group. In contrast, the mean lung fibre burden from 3 rats/dose group exposed to para-aramid fibres was 7.6×107 (±1.9×107) fibres per lung for the high dose group and 4.8×107 ( ±2.1×107) fibres/lung for the low dose group. In addition, the count median length of chrysotile fibres recovered from the lungs of exposed animals immediately after 2-week exposure was 3.5 µm, while the count median diameter was 0.15 µm. In contrast, the count median length of para-aramid fibres recovered from the lungs of exposed animals immediately after 2-week exposure was 8.6 µm, while the count median diameter was 0.3 µm (Figure 8.2(A) and (B); numerical data not shown). These data indicate that our attempts to size-separate Canadian chrysotile fibres were only partially successful. The lung burden data also suggest that comparisons of the effects of chrysotile vs paraaramid at high and low doses are difficult to make since the doses were not equivalent. Bronchoalveolar lavage data Two-week exposures to p-aramid fibrils or chrysotile asbestos fibres produced transient pulmonary inflammatory responses as measured by bronchoalveolar lavage fluid analysis (see Table 8.1). Light microscopic histopathology Exposures to p-aramid and chrysotile were associated with minimal to mild centriacinar inflammation and fibrosis (increased trichrome staining) immediately after and 5 days after 2-week exposures. Lesions were slightly more prominent in p-aramid-exposed rats due to increased inflammation. Lesions were less severe at 1 month and essentially resolved at 6 months

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Figure 8.1 (A) Chrysotile asbestos lengths—original generated sample for 4 different experiments. The graph depicts the fibre length distributions as assessed by scanning electron microscopy from four aerosol exposures prior to attempts to size separate the fibres. Fifty percent of the fibres from all four groups are less than 3–4 µm. (B) Distributions of size-separated chrysotile asbestos lengths used in the inhalation study from the high-dose

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Figure 8.1 Continued exposures and (C) the low-dose exposure groups. A casual glance at the two graphs B and C indicates that some success was attained in increasing the mean lengths in the aerosol of the generated chrysotile asbestos sample.

with only occasional centriacinar regions having slight, fibril-associated thickening of alveolar duct bifurcations. At 1 year postexposure, the lungs in p-aramid exposed rats were similar to controls. The 1-year chrysotileexposed animals are still in recovery. Pulmonary cell proliferation In chrysotile asbestos-exposed rats, substantial increases compared to controls in pulmonary cell proliferation indices were measured on terminal bronchiolar, parenchymal, visceral pleural/subpleural and mesothelial surfaces, and many of these effects were sustained through 3 months postexposure. These data demonstrate that 2-week chrysotile exposures produced a prolonged proliferative response in airway, alveolar and subpleural cells, as evidenced by the sustained effect through 3 months postexposure (Table 8.2). Pulmonary cell proliferation studies demonstrated that 2-week exposures to the high dose of p-aramid fibrils produced a transient increase in terminal bronchiolar and visceral pleural/subpleural cell labeling responses. No increases in lung parenchymal, or subpleural cell labeling indices were mea sured at any time period relative to sham controls. In addition, no

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Figure 8.2 (A) Scanning electron microscopy (SEM) micrograph of an aerosol filter containing a mixture of long and short chrysotile asbestos fibres (arrows). (B) An SEM micrograph of fibres recovered from the lung of a rat 3 months after 2-week chrysotile exposures. Note that most of the fibres are long (arrows), indicating that the long chrysotile asbestos fibres were retained in the lung while the shorter fibres were cleared from the respiratory tract.

increases in cell labeling indices were measured in animals exposed to a lower dose of p-aramid fibrils at any postexposure time period (Table 8.2).

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Table 8.1 Pulmonary inflammation and fibre biodurability in the lungs of chrysotile asbestos and p-aramid-exposed rats

0 h=immediately after exposure; 5 D=5 days; 1 M=1 month; 3 M=3 months; 6 M=6 months; ND=not determined.

Lung digestion/biodurability studies Preliminary dimensional analysis studies demonstrated that median lengths of fibres recovered from digested asbestos-exposed lung tissue were increased over time suggesting that short asbestos fibres were selectively cleared from the lungs, with apparent insignificant or pulmonary clearance and greater durability/retention of long fibres (Table 8.1). Preliminary studies with p-aramid fibrils recovered from the lungs of exposed rats are consistent with earlier data suggesting biodegradability of inhaled p-aramid fibrils (Warheit et al., 1992; Kelly et al., 1993) (Table 8.1). These data also are in agreement with the results of a current interim report authored by the Institute of Occupational Medicine in Edinburgh, Scotland. In addition, as previously reported (Warheit et al., 1992), a transient increase in fibre numbers at early postexposure time periods was measured following cessation of exposure. These results indicate that the increase in p-aramid fibres is due to fibre shortening and as a consequence, increased numbers of shorter fibres. This is accounted for by a substantial reduction in the median lengths of recovered fibres concomitant with only a slight decrease in fibre diameter.

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Table 8.2 Cell proliferation effects in chrysotile asbestos and p-Aramid-exposed rats

a p<0.05.

Discussion This multifunctional study was designed to compare the pulmonary toxic effects of inhaled, size-separated preparations of chrysotile asbestos fibres with p-aramid fibrils at similar aerosol fibre concentrations. Two-week exposures to p-aramid fibrils or chrysotile asbestos fibres produced transient pulmonary inflammatory responses as measured in fluids recovered from pulmonary lavage and this was consistent with histopathological assessments indicating that exposures produced an acute inflammatory response corresponding to alveolar thickening which peaked at 1 month post-exposure and became essentially reversible by 6 months postexposure. Cell proliferation studies with p-aramid fibrils demonstrated a transient increase in terminal bronchiolar and pleural/subpleural cell labeling immediately after exposure but not at 5 days or any other postexposure time period. In addition, no effects on cell turnover were measured at any postexposure time period in lung parenchymal, or mesothelial cells.

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The BrdU pulmonary cell labeling results demonstrating sustained proliferative effects in chrysotile-exposed rats presented here are consistent with findings from several other investigators (Brody and Overby, 1989; McGavran et al., 1990; Coin et al., 1992a). In studies by Brody and Overby (1989), acute inhalation exposures to chrysotile asbestos fibres produced a biphasic cell labeling response in the lungs of exposed rats and mice. This was characterized by dramatic increases in epithelial cell DNA synthesis, followed several days later by enhanced labeling of interstitial cells. In follow-up studies, a 3 day exposure prolonged the duration of increased cell labeling (Coin et al., 1992b). In another study, Coin et al., (1991) reported that a 5-h exposure to chrysotile fibres in mice produced substantial increases in mesothelial and subpleural cell labeling indices at 2 and 8 days postexposure. The finding of sustained subpleural and mesothelial cell proliferation in chrysotile-exposed rats was unexpected and raises the issue regarding the association of chrysotile with the development of mesothelioma. In this regard, inhalation of chrysotile asbestos fibres produced mesotheliomas in exposed rats (Wagner et al., 1974; Davis and Jones, 1988). The biodurability data reported here demonstrating retention or reduced clearance of long chrysotile fibres are consistent with the results of previous studies by Roggli and Brody (1984) and Bellmann et al., (1986, 1987). In contrast to the enhanced biodurability of chrysotile asbestos fibres, the results with p-aramid fibres suggest that the fibrils undergo biodegradability in the lungs of exposed rats. These findings confirm our earlier studies (Warheit et al., 1992) and are in concordance with the results of Kelly et al. (1993) and the recent findings of the IOM. In conclusion, size separation techniques for chrysotile asbestos fibres were partially successful in increasing median lengths from 3 µm to 6 µm. Histopathological studies demonstrated that both p-aramid and chrysotile produced a minimal to mild inflammatory response which produced thickening of the alveolar duct bifurcations. These effects peaked at 1 month postexposure and were essentially reversible by 6 months postexposure. Pulmonary cell labeling studies demonstrated substantial increases in lung parenchymal, airway, pleural/subpleural, and mesothelial cell proliferation effects following chrysotile exposures, suggesting that chrysotile produces a potent proliferative response in the airways, lung parenchyma, and subpleural/ pleural regions. In contrast, p-aramid exposures produced only transient effects in airway and subpleural regions. Fibre biopersistence/durability results thus far indicate that the long chrysotile fibres are retained in the lung or cleared at a slow rate. In contrast, p-aramid fibres have low biodurability in the lungs of exposed animals. In this regard, median lengths of chrysotile fibres recovered from

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exposed lung tissue were increased over time, while median lengths of paramid fibrils were decreased over time. It is concluded that the proliferative effects and enhanced biodurability of chrysotile that are associated with the induction of chronic disease do not occur with p-aramid fibrils. Therefore, inhalation of chrysotile asbestos fibres is likely to produce greater long-term pulmonary toxic effects in comparison to para-aramid fibrils. Acknowledgments This study was sponsored by the DuPont Co. and Akzo Nobel Corp. References BELLMANN, B., KONIG, H., MUHLE, H. and POTT, F., 1986, Chemical durability of asbestos and of man-made mineral fibres in vivo, J. Aerosol Sci., 17, 341–5. BELLMANN, B., MUHLE, H., POTT, F., KONIG, H., KLOPPEL, H. and SPURNY, K., 1987, Persistence of man-made mineral fibers (MMMF) and asbestos in rat lungs, Ann. Occup. Hyg., 31(4B), 693–709. BRODY, A.R. and OVERBY, L.H., 1989, Incorporation of tritiated thymidine by epithelial and interstitial cells in bronchiolar-alveolar regions of asbestosexposed rats, Am. J. Pathol., 134, 133–40. COIN, P.G., MOORE, L.B., ROGGLI, V. and BRODY, A.R., 1991, Pleural incorporation of 3H-TdR after inhalation of chrysotile asbestos in the mouse, Am. Rev. Respir. Dis., 143, A604. COIN, P.G., ROGGLI, V.L. and BRODY, A.R., 1992a, Deposition, clearance and translocation of chrysotile asbestos from peripheral and central regions of the rat lung, Environ, Res., 58, 97–116. COIN, P.G., ROGGLI, V. and BRODY, A.R., 1992b, Pulmonary fibrogenesis and BRDU incorporation after three consecutive inhalation exposures to chrysotile asbestos, Am. Rev. Respir. Dis., 145, A328. DAVIS, J.M.G. and JONES, A.D., 1988, Comparisons of the pathogenicity of long and short fibres of chrysotile asbestos in rats, Br. J. Exp. Pathol., 69, 717–37. DAVIS, J.M.G., ADDISON, J., BOLTON, R.E., DONALDSON, K. et al., 1986, The pathogenicity of long versus short fibre samples of amosite asbestos administered to rats by inhalation or intraperitoneal injection, Br. J. Exp. Pathol, 67, 415–30. KELLY, D.P., MERRIMAN, E.A., KENNEDY, G.L.JR. and LEE, K.P., 1993, Deposition, clearance, and shortening of Kevlar para-aramid fibrils in acute, subchronic, and chronic inhalation studies in rats, Fundam. Appl. Toxicol, 21, 345–54. LEE, K.P., KELLY, D.P., O’NEAL, F.O., STADLER, J.C. and KENNEDY, G. L.JR, 1988, Lung response to ultrafine Kevlar aramid synthetic fibrils following 2-year inhalation exposure in rats, Fundam. Appl. Toxicol., 11, 1– 20.

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McGAVRAN, P.D., BUTTERICK, C.J. and BRODY, A.R., 1990, Tritiated thymidine incorporation and the development of an interstitial lesion in the bronchiolar alveolar regions of the lungs of normal and complement deficient mice after inhalation of chrysotile asbestos, J. Environ. Pathol. Toxicol. Oncol, 9, 377–92. ROGGLI, V.L. and BRODY, A.R., 1984, Changes in numbers and dimensions of chrysotile asbestos fibers in lungs of rats following short-term exposure, Exp. Lung Res., 7, 133–47. WAGNER, J.C., BERRY, G., SKIDMORE, J.W. and TIMBRELL, V., 1974, The effects of the inhalation of asbestos in rats, Br. J. Cancer, 29, 252–70. WARHEIT, D.B., HILL, L.H. and BRODY, A.R., 1984a, Surface morphology and correlated phagocytic capacity of pulmonary macrophages lavaged from the lungs of rats, Exp. Lung Res., 6, 71–82. WARHEIT, D.B., CHANG, L.Y., HILL, L.H., HOOK, G.E.R., CRAPO, J.D. and BRODY, A.R., 1984b, Pulmonary macrophage accumulation and asbestosinduced lesions at sites of fiber deposition, Am. Rev. Respir. Dis., 129, 301. WARHEIT, D.B., CARAKOSTAS, M.C., HARTSKY, M.A. and HANSEN, J.F., 1991, Development of a short-term inhalation bioassay to assess pulmonary toxicity of inhaled particles: Comparisons of pulmonary responses to carbonyl iron and silica, Toxicol Appl. Pharmacol., 107, 350–68. WARHEIT, D.B., KELLAR, K.A. and HARTSKY, M.A., 1992, Pulmonary cellular effects in rats following aerosol exposures to ultrafine Kevlar® aramid fibrils: evidence for biodegradability of inhaled fibrils, Toxicol. Appl. Pharmacol, 116, 225– 39.

9 Pulmonary Hyperreactivity to Industrial Pollutants JÜRGEN PAULUHN Bayer AG, Wuppertal

Introduction Environmental agents, such as ozone, nitrogen dioxide, formaldehyde, and sulfur dioxide; occupational pollutants, including natural dusts (grain, red cedar, animal dander), irritant fumes or vapors, and organic acid anhydrides, reactive dyes, or (di)isocyanates can cause increases in airway reactivity. Airway hyperreactivity is defined as an exaggerated acute obstructive response of the airways to one or more nonspecific stimuli. The incriminated etiologic low-molecular-weight agents all share a common toxicological characteristic of being irritant in nature. In some cases, the agents are present as a gas, in others the inciting agent is an aerosol. As yet it is not clear, for instance, whether induced airway hyperreactivity is a dose-effect phenomenon and whether a brief high level exposure causes more prolonged or intense airways response. While the illness clinically simulates bronchial asthma and is associated with airway hyperreactivity, it is considered to be different from typical occupational asthma because of its rapid onset, specific relationship to a single environmental exposure, and no apparent preexisting period of sensitization to occur with the apparent lack of an allergic or immunologic etiology. Hence, this illness is termed reactive airways dysfunction syndrome, or RADS, because the characteristic finding is hyperreactivity of the airways (Brooks et al., 1985). Mechanisms to explain the development of RADS focus on the toxic effects of the irritant exposure on the airways. How this increased bronchial responsiveness is precisely triggered, amplified, sustained and how it relates to inflammatory events remains, to a certain extent, incompletely elucidated (Kay, 1991). A common pathologie accompaniment or cause of increased airway hyper-responsiveness is pulmonary inflammation. It is suggested that this inflammation is responsible for the change in histamine or cholinergic agonist responsiveness. Because subepithelial irritant receptors are superficial in location, they could be affected by an extensive bronchial inflammatory response which might occur after heavy irritant exposure.

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Subsequent re-epithelialization and probable reinervation of bronchial mucosa might drastically alter the threshold of the receptors and cause airways hyperreactivity. It has been hypothesized that damage to airway epithelium by irritant chemicals could decrease the threshold of sensory endings within the mucosa, resulting in increased afferent and efferent vagal activity. Airway mucosal inflammation, activation of airway afferent nerves, and the release of low-molecular-weight neuropeptides as mediators of inflammation are known to affect the tonus of the airway smooth muscle and may play a crucial role in the acute increase in airway hyperresponsiveness occurring after exposure to irritant or inflammatory stimuli. Additionally, inflammatory mediators may further attract and activate inflammatory cells, which themselves release a whole array of chemotactic and cytotoxic mediators that serve to perpetuate and amplify the inflammatory response. This complex interaction of different factors may result in epithelial desquamation, mucus gland hyperplasia, smooth muscle hypertrophy, and eventually render the airways hyperreactive to specific as well as nonspecific stimuli. Increased bronchial irritability, or hyperresponsiveness, to a wide variety of chemical agents and physical stimuli is also a major characteristic feature of bronchial asthma and the reactive airways dysfunction syndrome might clinically be indistinguishable from the asthma syndrome. Also for the latter, particular attention has been placed on the role of inflammation mediated influx of cells, mediator release and the interaction of irritant induced neurogenic and inflammatory factors. Neural control of airway caliber is far from being simple and it is likely to contribute to airway narrowing and bronchial hyper-responsiveness. Myelinated and nonmyelinated nerve fibers (C fibers) are involved in the sensory irritation response and their stimulation may result in release of specific neuropeptides, known to be potent releasers of mediators from airway mast cells (Barnes et al., 1991a, b; Nielsen, 1991). Specific neuropeptides are also known to attract eosinophils which can be stimulated to release cytotoxic mediators that may exacerbate these pseudoallergic-like responses even further. Experimental and clinical studies have intimated that there is reason to suspect that acute exposure to brief high-level concentrations of asthmagenic chemicals and the development of increased airway hyperresponsiveness are associated. Thus, it could be assumed that specific mast cell sensitization—in combination with neurogenic stimuli— amplify the inflammatory process and airway hyperresponsiveness. The corresponding increase in vagal activity would increase reflex release of acetylcholine and, correspondingly, may enhance airway responsiveness following the exogenous administration of cholinergic agents. Animal models of airway inflammation might allow us to investigate this relationship further. Models of allergic pulmonary inflammation have been developed in various animal species (Kips et al., 1992), using different

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method- ological approaches. In toxicology, the guinea-pig has been used for decades in order to evaluate the skin sensitizing properties of chemicals and proteins and has also been able to reproduce immediate-onset pulmonary hypersensitivity responses following inhalation of chemical haptens, their protein-conjugates or antigens. This animal model has therefore been used to disclose principles governing both the development of pulmonary hypersensitivity and airway hyperreactivity. Due to the guinea-pig’s abundant amount of smooth bronchial musculature, it is used as a physiologic elicitation model that reproduces bronchospasm upon challenge to specific or nonspecific stimuli. Other animal models designed to display many of the chronic features of hypersensitivity lung diseases characteristic of occupational asthma focus more on the induction of airway inflammation, the basic prerequisite for airway hyperreactivity. It should be noted, however, that the induction of asthma in the rat model, for example, commonly requires more aggressive protocols and more elaborate techniques to classify responses when compared with the guineapig elicitation model (vide infra). The guinea-pig model To date, practically all such models have relied upon the use of the guineapig, a species known to be sensitive for agents inducing bronchoconstriction and in which respiratory function and respiratory hypersensitivity can be measured readily. In addition, guinea-pigs are easy to handle, relatively inexpensive, and produce consistent bronchoconstrictive reactions. The models have utilized various modes of hapten or antigen administration and methods for detecting sensitization, It has been shown that guinea-pigs sensitized by inhalation exposure to either a free or a protein-bound chemical can be induced to exhibit changes in respiratory patterns following inhalation challenge with the same chemical in the free or in the form of its hapten-protein conjugate. In the guinea-pig no adjuvant is needed for successful lung sensitization. More recently it has been found that changes in sensitive respiratory parameters can also be provoked in dermally sensitized guinea-pigs by inhalation challenge with the free chemical or the hapten-protein conjugate (Botham et al., 1988; Pauluhn and Eben, 1991; Hayes et al., 1992). In attempting to derive an animal model that permits the identification of asthmagenic lowmolecularweight chemicals without the presence of overriding effects caused by toxic (irritant) airway inflammation the intradermal route of induction appears to be preferable. This route of induction also minimizes the risk of potential confounding effects attributable to irritant-induced nonspecific reactive airways dysfunction as a result of previous inhalation exposures (Briatico-Vangosa et al., 1993).

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For challenge exposures it appears to be advantageous to use the free chemical in slightly irritant concentrations rather than the proteinconjugate of the hapten. It is believed that the in vitro synthesis of the hapten-protein conjugate may not necessarily result in immunologically identical conjugates when compared with those produced under in vivo conditions. Also standardized procedures to synthesize and characterize hapten-protein conjugates of multifunctional, highly reactive chemicals are not yet established. On the other hand, an essential prerequisite for challenge exposures with the free chemical is the evaluation of the irritant threshold concentration of the hapten under investigation. The importance of concentration in distinguishing irritation from sensitization cannot be overstated and is one of the most critical determinants of this animal model. For volatile, irritant haptens the characteristic feature of upper respiratory tract irritation is the reflexively induced decrease in respiratory rate which is a common finding in laboratory rodents (Figure 9.1). Consistent with this approach, naive mice, rats and guinea-pigs were exposed for 45 min to slightly irritant concentrations of phenyl isocyanate (PI). As evident from Figure 9.1, the exposure to ca. 5 mg PI m−3 air provoked a decrease in respiratory rate of approximately 25–45%. The observation that remarkable differences in response patterns between mice, rats and guinea-pigs did not occur demon strate that irritant threshold concentrations obtained in mice may also be valid for guinea-pigs. Mainly for volatile chemicals attempts have been made to establish methods for the measurement and analysis of the irritant-induced changes in respiratory pattern in mice (Vijayaraghavan et al., 1993) and to understand the mechanisms of the irritant receptor stimulation (Nielsen, 1991). For volatile irritant haptens, such as PI, an unequivocal respiratory hypersensitivity response is characterized by a shallow rapid breathing pattern, i.e. a response opposite to that occurring as a result of upper respiratory tract irritation. For volatile irritant haptens this type of breathing pattern, however, can only be obtained when using the proteinconjugate of the hapten. The interpretation of changes in respiratory pattern induced by irritant particulates is less predictable because of the size-dependent deposition of particles within the respiratory tract. Irritant aerosols that evoke bronchial or pulmonary irritation may produce changes similar to those occurring following immediate-onset responses. Therefore, the selection of adequate haptenchallenge concentrations as well as the measurement of several breathing parameters is of primary importance. For such chemicals, currently the relative effectiveness of the acute high-concentration inhalation (single inhalation exposure of 15 min) and the high-dose intradermal route for sensitization of guinea-pigs had been investigated (Pauluhn and Eben, 1991; Pauluhn and Mohr, 1994). The airway function

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Figure 9.1 Time-response curves for respiratory rate from mice, rats and guineapigs during single 45-min exposures to appoximately 5 mg m−3 phenyl isocyanate. Data were normalized on pre-exposure values during 15-min air exposure. Data points for each concentration are the mean of four animals and were averaged for 45 s.

of conscious guinea-pigs that were sensitized to and challenged with 4,4′diphenylmethane-diisocyanate (MDI) aerosol or trimellitic anhydride (TMA) dust as well as their corresponding proteinconjugates was monitored plethysmographically. The airway hyper-responsiveness to subsequently increased inhaled acetylcholine (ACh) concentrations was assessed 1 day after the hapten challenge (Pauluhn, 1994). In most instances, selected morphological features of the airways (increased number of eosinophils in the bronchial mucosa and lung associated lymph nodes) were also taken into account. Collectively, it was noticed that elicitation of respiratory hypersensitivity is concentration-dependent and that challenge concentrations should slightly exceed the threshold concentration for irritation. The evaluation of eosinophils in subepithelial tissues and lung associated lymph nodes appears to provide an important independent adjunct to measurements of respiratory function. The combined assessment of specific pathologic features such as eosinophilic infiltration and the evaluations of several

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breathing parameters upon acetylcholine and hapten or conjugate challenge significantly enhance the diagnostic sensitivity of the guinea-pig model. From studies using single, brief high level aerosol or dust exposures for the induction of animals it can be concluded that previous high level exposures evoke bronchial hyperresponsiveness upon challenge at lower hapten concentrations when compared with intradermally sensitized animals. However, guinea-pigs sensitized intradermally to the volatile PI demonstrated remarkable immediate-type respiratory reactions only upon challenge with the conjugate and not with slightly irritant concentrations of the free PI. To study if phenyl isocyanate is capable of inducing a reactive airway or an asthma like syndrome, the subsequently described rat model was used. The rat model This animal model focuses on the induction of airway inflammation which comprises most of the characteristic features of asthma. It has been stated that respiratory hypersensitivity should depend on two separate factors: first, the degree of allergic airways, and second, the sensitivity to bronchoconstrictive mediators. Increasing evidence suggests that the eosinophils play a critical role in the pathogenesis of asthma and of other non-allergic hyperresponsive airway diseases. For the induction of the asthmatic state male rats were exposed for 2 consecutive weeks by inhalation (5 h day−1, 5 days week−1). The target concentrations of phenyl isocyanate were chosen on the basis of a single 45-min exposure study which suggested that approximately 1 mg m−3 air is the irritant threshold concentration for ‘any’ duration of exposure. The 2 week repeated inhalation study was designed to assess the functional, bio chemical and morphological signs of phenyl isocyanate induced lung disease and their regression during an observation period of approximately 2 months. The most characteristic features of asthma comprise an increased influx of eosinophilic granulocytes into the tissue of the airways, secretory cell hyper-plasia and metaplasia, smooth muscle hypertrophy and hyperplasia, epithelial desquamation, airway hyperresponsiveness, and eventually partial occlusion of the airway lumen with mucus and cellular debris. The formation of mucus plugs is a regular feature of asthma and accounts for most of the clinical, biochemical and physiological abnormalities. Histopathological evaluation of the respiratory tract indicated a bronchiolitis obliterans and smooth muscle hypertrophy in rats exposed to approximately 7 mg m−3 air, whereas only minimal effects were found following 4 mg m−3 air. Lung function measurements revealed that some rats were hyperresponsive to an ACh-stimulus. As shown in Figure 9.2, also the increase in shunt blood (Qs/Qt) anddecrease in forced expiratory flow rates (MMEF) as well as mucus products (sialomucins), polymorphonuclear

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Figure 9.2 Relative comparison of sensitive diagnostic parameters in rats exposed to either 0 (air), 1, 4 and 7 mg PI m−3 air for 2 consecutive weeks (6 h day−1, 5 days week−1). The measurements were performed in weeks 3 and 9. Abbreviations: MMEF: Maximal mid-expiratory flow rate, Qs/Qt: venous admixture, PMN: polymorphonuclear cells, Eos: eosinophilic granulocytes, Sialomucins: total sialic acid (after hydrolysis).

cells, including eosinophils, in the bronchoalveolar lavage fluid (BALF) were consistent with an asthma like syndrome. As evident from Figure 9.2, the changes observed in rats exposed to 7 mg m−3 air did not fully regress during an observation period of approximately 2 months. Conclusion Experimental evidence suggests that changes within the respiratory tract leading to the reactive airway dysfunction syndrome and/or asthma are fully consistent with an inflammatory response involving tissue of direct contact. The toxicity of irritant chemicals known to induce such illness is highly focal, and the variability of response in different regions of the respiratory tract could be a result of the actual concentration of the toxicant reaching various airway levels. Determination of immunologic etiology is particularly important for chemical allergy since all recognized low-molecular-weight chemical sensitizers are also respiratory irritants and in sufficient concentrations can cause airway constriction by nonimmunological mechanisms. As shown by studies using phenyl

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isocyanate, damage of the airways is characterized by a steep concentrationresponse curve. Based on acute 45-min exposure of rats the threshold for respiratory tract irritation is approximately 1 mg m−3. Exposures equaling this concentration were tolerated without exposure-related effects, whether exposure occurred singly for 45-min or repeatedly for 2 weeks. Marginal effects were observed at 4 mg m−3, all effects, including mortality, were produced at 7 mg m−3. This demonstrates that selection of appropriate exposure concentrations appears to be most critical in the rat model. The assessment of diagnostic sensitivity of the methods used to probe damage to the respiratory tract demonstrated that respiratory function data, blood gas measurements, and BALF analysis facilitate a meaningful interpretation of the effects observed and are important adjuncts to common inhalation toxicological studies on rats to describe quantitatively the diseased state of the lung. The guinea-pig model is experimentally less demanding and therefore can suitably be used as a screening test for respiratory sensitization, as far as the limitations of this model are taken into account. Studies on guinea-pigs demonstrate that elicitation of respiratory hypersensitivity is challengeconcentration dependent and that the concentrations used should slightly exceed the threshold concentration for irritation to maximize the magnitude of the response. However, sensitization by inhalation may increase the susceptibility to irritant stimuli and thus confounds the selection of the most appropriate concentration for challenge. The combined approach of evaluating several breathing parameters, e.g. respiratory rate, flow- and volume-derived parameters, during both the hapten (free or conjugated) and the ACh challenge provides a promising method to distinguish specific and nonspecific hypersensitivity responses. Furthermore, it is critically important to assess the respiratory irritant potency of the compound under investigation. For potent irritant substances such as volatile isocyanates, challenge with the haptenprotein conjugate minimizes the likelihood to confound specific hypersensitivity responses with those evoked merely by irritation. Taking all imponderable factors into consideration, it appears that the guinea-pig intradermalinduction inhalation-challenge protocol is adequately susceptible to identify potent respiratory tract sensitizers. However, if the airway inflammation related features of asthma are the endpoints of primary interest other animal models appear to be more appropriate. References BARNES, P.J., BARANIUK, J.N. and BELVISI, M.G., 1991a, Neuropeptides in the respiratory tract (Part II). Am. Rev. Respir. Dis., 144, 1391–9.

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BARNES, P.J., CHUNG, K.F., PAGE, C.P., 1991b, Pharmacology of asthma, Chapter 3, Inflammatory Mediators in Page, C.P. and Barnes, P.J. (Eds.), pp 54–106. Handbook of Experimental Pharmacology, Berlin, Heidelberg, New York: Springer-Verlag. BOTHAM, P.A., HEXT, P.M., RATTRAY, N.J., WALSH, S.T. and WOODCOCK, D.R., 1988, Sensitisation of guinea-pigs by inhalation exposure to lowmolecular-weight chemicals, Toxicol. Lett., 41, 159–73. BRIATICO-VANGOSA, G, BRAUN, C.J.L., COOKMAN, G., HOFMANN, T., KIMBER, I., LOVELESS, S.E., MORROW, T., PAULUHN, J., SØRENSEN, T. and NIESSEN, H.J., 1993, Respiratory Allergy. ECETOC Monograph No. 19. BROOKS, S.M., WEISS, M.A. and BERNSTEIN, I.L., 1985, Reactive airway dysfunction syndrome (RADS). Persistent asthma syndrome after high level irritant exposure, Chest, 88, 376–84. HAYES, J.P., DANIEL, H.R., TEE, R.D., BARNES, P.J., NEWMAN-TAYLOR, A.J. and CHUNG, K.F., 1992, Bronchial hyperreactivity after inhalation of trimellitic anhydride dust in guinea-pigs after intradermal sensitization to the free hapten, Am. Rev. Respir. Dis., 146, 1311–14. KAY, A.B., 1991, Asthma and inflammation, J. Allergy Clin. Immunol., 87, 893– 910. KIPS, J.C., CUVELIER, C.A. and PAUWELS, R.A., 1992, Effect of acute and chronic antigen inhalation on airway morphology and responsiveness in actively sensitized rats, Am. Rev. Respir. Dis., 145, 1306–10. NIELSEN, G.D., 1991, Mechanisms of activation of the sensory irritant receptor by airborne chemicals, Crit. Rev. Toxicol, 21, 183–208. PAULUHN, J., 1994, Test methods for respiratory sensitization in use of mechanistic information in risk assessment, EUROTOX Proceedings, Arch. Toxicol., suppl. 16, 77–86. PAULUHN, J. and EBEN, A., 1991, Validation of a non-invasive technique to assess immediate or delayed onset of airway hypersensitivity in guinea-pigs, J. Appl. Toxicol, 11, 423–31. PAULUHN, J. and MOHR, U., 1994, Assessment of respiratory hypersensitivity in guinea-pigs sensitized to diphenylmethane-4,4'-diisocyanate (MDI) and challenged with MDI, acetylcholine or MDI-albumin conjugate, Toxicology (in press). VIJAYARAGHAVAN, R., SCHAPER, M., THOMPSON, R., STOCK, M.F. and ALARIE, Y., 1993, Characteristic modifications of the breathing pattern of mice to evaluate the effects of airborne chemicals on the respiratory tract, Arch. Toxicol, 67, 478–90.

10 Mechanisms of Pulmonary Sensitization IAN KIMBER Zeneca Central Toxicology Laboratory, Macclesfield

Introduction A wide range of chemicals is known to cause allergic contact dermatitis. It is apparent, however, that chemicals also have the potential to provoke other forms of allergy and of growing concern is pulmonary sensitization. Examples of chemicals identified as human respiratory allergens are listed in Table 10.1. Respiratory allergic hypersensitivity is characterized by pulmonary reactions which occur normally in only a proportion, and frequently in only a small proportion, of exposed individuals. In those who are sensitized, respiratory reactions can be provoked by atmospheric concentrations of the causative chemical allergen which were tolerated previously and which are without effect in the non-sensitized population (Newman Taylor, 1988). Almost invariably there is a latent period between the onset of exposure and the development of respiratory symptoms such as asthma and rhinitis. By definition, allergy, including sensitization of the respiratory tract, results from the stimulation of specific immune responses by the causative agent. Although it is assumed frequently that effective allergic sensitization of the respiratory tract results largely or wholly from inhalation exposure, this is not necessarily the case. Allergic reactions manifest in a particular organ commonly result from the local provocation by the inducing agent of a systemically sensitized individual. There is no reason to suppose that the quality of immune response necessary for sensitization of the respiratory tract may not result from exposure to the chemical allergen at a different site. Consistent with this is evidence that occupational respiratory allergy may be caused by dermal contact with the chemical (Karol, 1986; Nemery and Lenaerts, 1993). Furthermore, it has been reported that respiratory rate changes can be provoked by inhalation exposure of guinea pigs sensitized previously by either topical or subcutaneous treatment with the same chemical (Karol et al., 1981; Rattray et al., 1994). Despite the fact that, in practice, pulmonary sensitization may not be caused exclusively by

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Table 10.1 Chemicals identified as human respiratory allergens

inhalation of the chemical allergen, it is likely that this is an important route of exposure in the occupational setting. It is well established that respiratory sensitization caused by protein aeroallergens is effected by IgE antibody. This class of antibody in man is homocytotropic and is able to associate, via specific membrane receptors, with mast cells, including mast cells in the respiratory tract. Following subsequent exposure of the sensitized individual to the same allergen, mast cell-bound IgE is cross-linked and this, in turn, results in mast cell degranulation and the release of both preformed and newly-synthesized mediators which provoke acute inflammatory reactions. In the case of sensitization of the respiratory tract caused by chemicals, however, an invariable association with the presence of specific IgE antibody has failed to emerge. Although IgE antibody specific for all recognized chemical respiratory allergens has been demonstrated, it is not uncommonly the case that individuals displaying symptoms of pulmonary hypersensitivity have been reported to lack demonstrable IgE. This may suggest that immunological processes independent of IgE antibody may play a decisive role in the induction of respiratory sensitization. An alternative explanation is that inappropriate or insensitive techniques have been employed for serological analysis and that IgE antibody may be associated more commonly than suspected previously with chemical respiratory allergy. In this context it is relevant that it has been found that positive skin prick tests can be provoked in patients sensitized to acid anhydrides who, on the basis of radioallergosorbent tests (RAST), were found to lack measurable levels of serum IgE antibody (Drexler et al., 1993). Despite the absence of formal confirmation that there exists a universal causal relationship between specific IgE and pulmonary hypersensitivity induced by chemicals,

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it remains likely that this class of antibody is responsible, in at least the majority of cases, for the acute onset symptoms associated with respiratory allergy (Karol et al., 1994). The induction and regulation of IgE responses IgE antibody responses are subject to a variety of immunoregulatory control mechanisms. Chief among these are the stimulatory and inhibitory actions of cytokines which serve to influence the induction and duration of IgE responses. It has been found in mice that interleukin 4 (IL-4) is necessary for the initiation and maintenance of IgE antibody production (Finkelman et al., 1988b). The essential role for this cytokine in IgE responses has been emphasized further by studies of mice homozygous for a mutation that inactivates the gene for IL-4. These animals lack detectable serum IgE and fail to mount IgE responses (Kuhn et al., 1991). Importantly, in mice which produce constitutively high levels of IL-4, significantly elevated concentrations of serum IgE are evident (Burstein et al., 1991). A balance to the promotional influence of IL-4 is provided by interferon (IFN- ), a cytokine which exerts an inhibitory affect on IgE responses (Finkelman et al., 1988a). The reciprocal antagonistic activity of these cytokines is not restricted to the mouse, IL-4 and IFN- have been found to regulate human IgE production (Del Prete et al., 1988; Pene et al., 1988). The cytokines which influence the integrity of IgE responses are the products of discrete subpopulations of T helper (Th) cells, lymphocytes characterized by possession of the CD4 membrane determinant. It has been found in both mouse and man that there exists a functional heterogeneity among Th cells. Two major populations, designated Th1 and Th2, have been described (Mosmann and Coffman, 1989; Romagnani, 1991). It is believed currently that these subsets represent the most differentiated forms of Th cells and develop from less mature precursors as the immune response evolves (Mosmann et al., 1991). The major functional distinction between Th1 and Th2 cells resides in the spectrum of cytokines which they elaborate (Mosmann and Coffman, 1989). The cytokine products of murine Th1 and Th2 cells are displayed in Table 10.2. It has been reported previously that chemicals known to cause respiratory hypersensitivity in man induce in mice immune responses characteristic of Th2 cell activation, stimulate the production of specific IgE antibody and cause an increase in the serum concentration of IgE. Conversely, chemical allergens considered not to cause respiratory sensitivity, but which are nevertheless able to induce skin sensitization, elicit instead Th1-type responses. In the latter case, immune responses are characterized by comparatively high levels of IgG2a antibody (an isotype known to be upregulated by IFN- ) and the absence of specific IgE (Dearman and Kimber, 1991, 1992; Dearman et al., 1991, 1992a,c,d,

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Table 10.2 The cytokine products of murine Th1 and Th2 cells

From: Mosmann and Coffman (1989).

1994). The implication is that certain chemicals favour the development of Th2 cells which will then synthesize and secrete IL-4 and thereby encourage IgE antibody responses and mast cell sensitization. The converse is that other classes of chemical allergen preferentially stimulate Th1 cells and IFNproduction. Such conditions will be nonpermissive for IgE antibody production and cell-mediated immune responses, including contact sensitization, will be favoured instead. A selective stimulation by different classes of chemical sensitizers of divergent Th responses may provide an explanation at the cellular level for the observation that chemicals vary with respect to the nature of allergic reactions that they will elicit preferentially in man. The stimulation by chemical allergens of differentiated Th cell responses may have implications for allergic disease other than the regulation of IgE antibody. It is known for instance that IL-3, IL-4 and IL-10, all of which are products of murine Th2 cells (Table 10.2), act as mast cell growth factors or cofactors (Smith and Rennick, 1986; Thompson-Snipes et al., 1991). Moreover, IL-5 is a growth and differentiation factor for eosinophils (Yokota et al., 1987) and serves to regulate the accumulation of these cells at the site of allergeninduced hypersensitivity reactions in the respiratory tract (Gulbenkian et al., 1992). It has been found recently that the cytokines IL-3 and IL-4 also enhance the secretory activity of mast cells following activation (Coleman et al., 1993). Antagonistic and inhibitory influences of Th cell products may also affect the elicitation of allergic reactions. It has been found that IFN- not only suppresses the secretory function of mast cells (Holliday et al., 1994), but also antagonizes the antigen-induced infiltration of eosinophils into the respiratory tract of sensitized mice (Iwamoto et al., 1993). Contact allergic reactions may in theory be regulated by Th2 cytokines. It has been shown that IL-4 and IL-10 act in concert to inhibit Th1 cell function and to

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depress cell-mediated immunity (Powrie et al., 1993) and that Il–4 is able to reduce significantly the severity of contact allergic reactions in mice (Gautam et al., 1992). Taken together the available data suggest that the selective stimulation of Th cell responses and the consequent balance created between Th1- and Th2-derived cytokines will have an important impact on both the induction and elicitation stages of allergy. It is perhaps not surprising, therefore, that there is increasing evidence for selective Th responses in human allergic disease. Clones of T lymphocytes specific for aeroallergens such as house dust mite and grass pollen, which cause IgE-mediated respiratory allergic reactions in susceptible individuals, have been shown to elaborate Th2 cytokines, but not IFN- (Parronchi et al., 1991). A predominance of the Th2-type cells has been found at sites of skin reactions in atopic individuals (Kay et al., 1991) and increased numbers of IL-4+ T lymphocytes have been identified in the nasal mucosa in allergen-induced rhinitis (Ying et al., 1994). By contrast, human immune responses to nickel, a common cause of allergic contact dermatitis, are characterized by the selective activation of Th1-type cells. Allergen-specific T lymphocyte clones isolated from the peripheral blood of patients sensitized to nickel have been found to secrete only low or undetectable amounts of IL-4 and IL-5, but high levels of IFN(Kapsenberg et al., 1991). Although the relative contribution of Th1 and Th2 cells during immune responses, and in particular the relative availability of IL-4 and IFN- , is likely to play a predominant role in the regulation of IgE antibody, other factors may be relevant. Not least, the priming of Th1 cells for the production of IFN- may in turn be dependent upon the action of another cytokine, interleukin 12 (IL-12) (Manetti et al., 1994; Morris et al., 1994; Schmitt et al., 1994). It has been demonstrated also that CD8+ T lymphocytes exert an important immunoregulatory influence on IgE responses (Kemeny et al., 1994; Renz et al., 1994), possibly via downregulation of CD4+ Th2 cell development (Noble et al., 1993). It is clear that conditions outwith the immune system also influence the magnitude of IgE responses. Certainly genetic predisposition is an important, although poorly understood factor. In addition, there have been suggestions that cigarette smoking and exposure to certain environmental pollutants may result in increased IgE levels and may also serve to aggravate asthma (Zetterstrom et al., 1981; Muranka et al., 1986; Wardlaw, 1993). Cell-mediated immune responses in chemical respiratory allergy The elicitation of chemical respiratory hypersensitivity may be associated with both immediate-onset and late phase reactions. While IgE antibody

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and local degranulation of mast cells may be necessary for acute symptoms, late asthmatic responses appearing some hours following exposure are characterized by an infiltration of mononuclear cells and increased numbers of leucocytes in bronchoalveolar lavage fluid. Chronic inflammation is an important component of asthma and, in addition to mononuclear cell accumulation, is characterized by mucus production, the destruction and sloughing of airway epithelial cells and subepithelial fibrosis secondary to collagen deposition. Eosinophils, acting together with infiltrating T lymphocytes, play a pivotal role in chronic bronchial inflammation (Corrigan and Kay, 1992). It is apparent also that the generation of eosinophilia in the respiratory tract is influenced markedly by Th cell products. As described previously, IL-5 effects the accumulation of eosinophils at the site of hypersensitivity reactions in respiratory tissues, while IFN- , secondary to an inhibition of CD4+ cell infiltration, antagonizes this process (Gulbenkian et al., 1992; Iwamoto et al., 1993). It may prove that the cell-mediated immune processes relevant to the development of respiratory hypersensitivity and asthma are also a function of Th cell heterogeneity. Certainly the stimulation of Th2 cell activation will have profound effects on all stages of respiratory allergy. The infiltration of such cells into sites of encounter with inducing allergen, a process perhaps facilitated by vasodilation resulting from mast cell degranulation, will provide a local source of cytokines such as IL-4 and IL-5. Mast cell secretory activity will be potentiated by the former and eosinophil accumulation triggered by the latter. That Th2 cells do in fact accumulate in the area of immediate-type hypersensitivity reactions is supported by the studies of Kay et al. (1991) who demonstrated that the cells infiltrating lesional skin at the sites of late phase cutaneous reactions in atopic patients produce IL-3, IL-4, IL-5 and GM-CSF, but not IFN- . Practical applications In the course of investigations designed to examine the characteristics of immune responses induced in mice by chemical sensitizers it was found that only those materials known to cause respiratory hypersensitivity in man provoked in mice a substantial increase in the serum concentration of IgE; a phenomenon thought to reflect the selective stimulation of Th2 celltype responses by this class of allergen. It was observed also that contact allergens known or suspected not to cause occupational respiratory hypersensitivity failed to result in similar changes in serum IgE levels (Dearman and Kimber, 1991, 1992; Dearman et al., 1992a,d). The differential ability of chemical respiratory and contact allergens to stimulate changes in the concentration of serum IgE in mice forms the basis of a novel approach to the identification of chemicals which have the potential to cause sensitization of the respiratory tract. This method, the

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mouse IgE test (Dearman et al., 1992b, Kimber and Dearman, 1993) is being evaluated currently in the context of internal and inter-laboratory validation studies. References BERNSTEIN, D.I., PATTERSON, R. and ZEISS, C.R., 1982, Clinical and immunological evaluation of trimellitic anhydride- and phthalic anhydrideexposed workers using a questionnaire and comparative analysis of enzymelinked immunosorbent and radioimmunoassay studies, Journal of Allergy and Clinical Immunology, 69, 311–18. BURSTEIN, H.J., TEPPER, R.I., LEDER, P. and ABBAS, A.K., 1991, Humoral immune function in IL-4 transgenic mice, Journal of Immunology, 147, 2950– 6. CARTIER, A., CHAN H., MALO, J-L., PINEAU, L., TSE, K.S. and CHANYEUNG, M., 1986, Occupational asthma caused by eastern white cedar (Thuja occidentalis) with demonstration that plicatic acid is present in this wood dust and is the causal agent, Journal of Allergy and Clinical Immunology, 77, 639–45. COLEMAN, J.W., HOLLIDAY, M.R., KIMBER, I., ZSEBO, K.M. and GALLI, S. J., 1993, Regulation of mouse peritoneal mast cell secretory function by stem cell factor, IL-3 or IL-4, Journal of Immunology, 150, 556–62. CORRIGAN, C.J. and KAY, A.B., 1992, T cells and eosinophils in the pathogenesis of asthma, Immunology Today, 13, 501–7. DEARMAN, R.J. and KIMBER, I., 1991, Differential stimulation of immune function by respiratory and contact chemical allergens, Immunology, 72, 563– 70. DEARMAN, R.J. and KIMBER, I., 1992, Divergent immune responses to respiratory and contact chemical allergens: antibody elicited by phthalic anhydride and oxazolone, Clinical and Experimental Allergy, 22, 241–50. DEARMAN, R.J., HEGARTY, J.M. and KIMBER, I., 1991, Inhalation exposure of mice to trimellitic anhydride induces both IgG and IgE anti-hapten antibody, International Archives of Allergy and Applied Immunology, 95, 70–6. DEARMAN, R.J., BASKETTER, D.A., COLEMAN, J.W. and KIMBER, I., 1992a, The cellular and molecular basis for divergent allergic responses to chemicals, Chemical-Biological Interactions, 84, 1–10. DEARMAN, R.J., BASKETTER, D.A. and KIMBER, I., 1992b, Variable effects of chemical allergens on serum IgE concentrations in mice. Preliminary evaluation of a novel approach to the identification of respiratory sensitizers, Journal of Applied Toxicology, 12, 317–23. DEARMAN, R.J., MITCHELL, J.A., BASKETTER, D.A. and KIMBER, I., 1992c, Differential ability of occupational chemical contact and respiratory allergens to cause immediate and delayed dermal hypersensitivity reactions in mice, International Archives of Allergy and Immunology, 97, 315–21.

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DEARMAN, R.J., SPENCE, L.M. and KIMBER, I., 1992d, Characterization of murine immune responses to allergenic diisocyanates, Toxicology and Applied Pharmacology, 112, 190–7. DEARMAN, R.J., RAMDIN, L.S.P., BASKETTER, D.A. and KIMBER, I. 1994, Inducible interleukin-4-secreting cells provoked in mice during chemical sensitization, Immunology, 81, 551–7. DEL PRETE, G., MAGGI, E., PARRONCHI, P., CHRETIEN, I., TIRI, D., MACCHIA, D., RICI, M., ANSARI, A.A. and ROMAGNANI, S., 1988, IL-4 is an essential factor for the IgE synthesis induced in vitro by human T cell clones and their supernatants, Journal of Immunology, 140, 4193–8. DOCKER, A., WATTIE, J.M., TOPPING, M.D., LUCZYNSKA, C.M., NEWMAN TAYLOR, A.J., PICKERING, C.A.C., THOMAS, P. and GOMPERTZ, D., 1987, Clinical and immunological investigations of respiratory disease in workers using reactive dyes, British Journal of lndustrial Medicine, 44, 534–41. DREXLER, H., SCHALLER, K-H., WEBER, A., LETZEL, S. and LEHNERT, G., 1993, Skin prick tests with solutions of acid anhydrides in acetone, International Archives of Allergy and Immunology, 100, 251–5. FINKELMAN, F.D., KATONA, I.M., MOSMANN, T.R. and COFFMAN, R.L., 1988a, IFN- regulates the isotypes of Ig secreted during in vivo humoral immune responses, Journal of Immunology, 140, 1022–7. FINKELMAN, F.D., KATONA, I.M., URBAN, J.F.JR, HOLMES, J., OHARA. J., TUNG, A.S., SAMPLE, J.G. and PAUL, W.E., 1988b, IL-4 is required to generate and sustain in vivo IgE responses, Journal of Immunology, 141, 2335– 41. GAUTAM, S.C., CHIKKALA, N.F. and HAMILTON, T.A., 1992, Antiinflammatory action of IL-4. Negative regulation of contact sensitivity to trinitrochlorobenzene, Journal of Immunology, 148, 1411–15. GULBENKIAN, A.R., EGAN, R.W., FERNANDEZ, X., JONES, H., KREUTNER, W., KUNG, T., PAYVANDI, F., SULLIVAN, L., ZURCHER, J.A. and WATNIK, A.S., 1992, Interleukin-5 modulates eosinophil accumulation in allergic guinea pig lung, American Review of Respiratory Diseases, 146, 263–9. HOLLIDAY, M.R., BANKS, E.M. S., DEARMAN, R.J., KIMBER, I. and COLEMAN, J.W., 1994. Interactions of IFN- with IL-3 and IL-4 in the regulation of serotonin and arachidonate release from peritoneal mast cells, Immunology, 82, 70–4. HOWE, W., VENABLES, K.M., TOPPING, M.D., DALLY, M.B., HAWKINS, R., LAW, J.S. and NEWMAN TAYLOR, A.J., 1983, Tetrachlorophthalic anhydride asthma: evidence for specific IgE antibody, Journal of Allergy and Clinical Immunology, 7l, 5–11. IWAMOTO, I., NAKAJIMA, H., ENDO, H. and YOSHIDA, S., 1993, Interferon regulates antigen-induced eosinophil recruitment into the mouse airways by inhibiting infiltration of CD4+ T cells, Journal of Experimental Medicine, 177, 573–6. KAPSENBERG, M.L., WIERENGA, E.A., Bos, J.D. and JANSEN, H.M., 1991, Functional subsets of allergen-reactive human CD4+ T cells, Immunology Today, 12, 392–5. KAROL, M.H., 1986, Respiratory effects of inhaled isocyanates, CRC Critical Reviews in Toxicology, 16, 349–79.

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KAROL, M.H., HAUTH, B.A., RILEY, E.J. and MAGRENI, C.M., 1981, Dermal contact with toluene diisocyanate (TDI) produces respiratory tract hypersensitivity in guinea pigs, Toxicology and Applied Pharmacology, 58, 221–30. KAROL, M.H., TOLLERUD, D.J., CAMPBELL, T.P., FABBRI, L., MAESTRELLI, P., SAETTA, M. and MAPP, C.E., 1994, Predictive value of airways hyperresponsiveness and circulating IgE for identifying types of responses to toluene diisocyanate inhalation challenge, American Journal of Respiratory and Critical Care Medicine, 149, 611–15. KAY, A.B., YING, S., VARNEY, V., GAGA, M., DURHAM, S.R., MOQBEL, R., WARDLAW, A.J. and HAMID, Q., 1991, Messenger RNA expression of the cytokine gene cluster, interleukin 3 (IL-3), IL-4, IL-5 and granulocyte/ macrophage colony stimulating factor, in allergen-induced last phase reactions in atopic subjects, Journal of Experimental Medicine, 173, 775–8. KEMENY, D.M., NOBLE, A., HOLMES, B.J. and DIAZ-SANCHEZ, D., 1994, Immune regulation: a new role for the CD8+ T cell, Immunology Today, 15, 107–10. KIMBER, I. and DEARMAN, R.J., 1993, Approaches to the identification and classification of chemical allergens in mice, Journal of Pharmacological and Toxicological Methods, 29, 11–16. KUHN, R., RAJEWSKY, K. and MULLER, W., 1991, Generation and analysis of interleukin-4 deficient mice, Science, 254, 707–10. MACCIA, C.A. BERNSTEIN, I.L., EMMETT, E.A. and BROOKS, S.M. 1976, In vitro demonstration of specific IgE in phthalic anhydride sensitivity, American Review of Respiratory Disease, 113, 701–4. MANETTI, R., GEROSA, F., GIUDIZI, M.G., BIAGIOTTI, R., PARRONCHI, P., PICCINNI, M-P., SAMPOGNARO, S., MAGGI, E., ROMAGNANI, S. and TRI NCHIERI, G., 1994, Interleukin 12 induces stable priming for interferon (IFN- ) production during differentiation of human T helper (Th) cells and transient IFN- production in established Th2 cell clones, Journal of Experimental Medicine, 179, 1273–83. MOLLER, D.R., GALLAGHER, J.S., BERNSTEIN, D.I., WILCOX, T.G., BURROUGHS, H.E. and BERNSTEIN, I.L., 1985, Detection of IgE-mediated respiratory sensitization in workers exposed to hexahydrophthalic anhydride, Journal of Allergy and Clinical Immunology, 15, 663–72. MORRIS, S.C., MADDEN, K.B., ADAMOVICZ, J.J., GAUSE, W.C., HUBBARD, B.R., GATELY, M.K. and FINKELMAN, F.D., 1994, Effects of IL-12 on in vivo cytokine gene expression and Ig isotype selection, Journal of Immunology, 152, 1047–56. MOSMAN, T.R. and COFFMAN, R.L., 1989, Heterogeneity of cytokine secretion patterns and functions of helper T cells, Advances in Immunology, 46, 111– 47. MOSMANN, T.R., SCHUMACHER, J.H., STREET, N.F., BUDD, R., O’GARRA, A., FONG, T.A.T., BOND, M.W., MOORE, K.W.M., SHER, A. and FIORENTINO, D.F., 1991, Diversity of cytokine synthesis and function of mouse CD4+ T cells, Immunological Reviews, 123, 209–29. MURANKA, M., SUZUKI, S., KOIZUMI, K., TAKAFUJI, S., MIYAMOTO, T., IKEMURI, R. and TOKIWA, H., 1986, Adjuvant activity of diesel-exhaust

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particulates for the production of IgE antibody in mice, Journal of Allergy and Clinical Immunology, 77, 616–23. MURDOCH, R.D., PEPYS, J. and HUGHES, E.G., 1986. IgE antibody responses to platinum group metals: a large scale refinery survey, British Journal of Industrial Medicine, 43, 37–43. NEMERY, B. and LENAERTS, L., 1993, Exposure to methylene diphenyl diisocyanate in coal mines, Lancet, 341, 318. NEWMAN TAYLOR, A.J., 1988, Occupational asthma, Postgraduate Medical Journal, 64, 505–10. NOBLE, A., STAYNOV, D.Z., DIAZ-SANCHEZ, D., LEE, T.J. and KEMENY, D. M., 1993, Elimination of IgE regulatory rat CD8+ T cells in vivo increases the co-ordinate expression of Th2 cytokines IL-4, IL-5 and IL-10, Immunology, 80, 326–9. O’BRIEN, I.M., HARRIES, M.G., BURGE, P.S. and PEPYS, J., 1979, Toulene diisocyanate-induced asthma. 1. Reactions to TDI, MDI, HDI and histamine, Clinical Allergy, 9, 1–6. PARRONCHI, P., MACCHIA, D., PICCINI, M-P., BISWAS, P., SIMONELLI, C., MAGGI, E., RICCI, M., ANSARI, A.A. and ROMAGNANI, S., 1991, Allergen and bacterial antigen-specific T cell clones established from atopic donors show a different profile of cytokine production, Proceedings of the National Academy of Sciences USA, 88, 4538–42. PENE, J., ROUSSET, F., BRIERE, F., CHRETIEN, I., PALIARD, X., BANCHEREAU, J., SPITS, H. and DE VRIES, J.E., 1988, IgE production by normal human B cells induced by alloreactive T cell clones is mediated by 11–4 and suppressed by IFN- , Journal of Immunology, 141, 1218–24. POWRIE, F., MENON, S. and COFFMAN, R.L., 1993, Interleukin-4 and interleukin-10 synergize to inhibit cell-mediated immunity in vivo, European Journal of Immunology, 23, 3043–9. QUIRCE, S., CUEVAS, M., OLAGUIBEL, J.M. and TABAR, A.I., 1994, Occupa tional asthma and immunological responses induced by inhaled carmine among employees at a factory making natural dyes, Journal of Allergy and Clinical Immunology, 93, 44–52. RATTRAY, N.J., BOTHAM, P.A., HEXT, P.M., WOODCOCK, D.R., FIELDING, I., DEARMAN, R.J. and KIMBER, I., 1994, Induction of respiratory hypersensitivity to diphenylmethane-4,4′-diisocyanate (MDI) in guinea pigs. Influence of route of exposure, Toxicology, 88, 15–30. RENZ, H., LACK, G., SALOGA, J., SCHWINZER, R., BRADLEY, K., LOADER, J., KUPFER, A., LARSEN, G.L. and GELFAND, E.W., 1994, Inhibition of IgE production and normalization of airways responsiveness by sensitized CD8 T cells in a mouse model of allergen-induced sensitization, Journal of Immunology, 152, 351–60. ROMAGNANI, S., 1991, Human TH1 and TH2 subsets: doubt no more, Immunology Today, 12, 256–7. SCHMITT, E., HOEHN, P., HUELS, C., GOEDERT, S., PALM, N., RUDE, E. and GERMANN, T., 1994, T helper type 1 development of naive CD4+ T cells requires the coordinate action of interleukin-12 and interferon- and is inhibited by transforming growth factor-β, European Journal of Immunology, 24, 793–8.

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SMITH, C.A. and RENNICK, D.M., 1986, Characterization of a murine lymphokine distinct from interleukin 2 and interleukin 3 (IL-3) possessing a Tcell growth factor and mast cell growth factor activity that synergizes with IL-3, Proceedings of the National Academy of Sciences USA, 83, 1857–61. THOMPSON-SNIPES, L., DHAR, V., BOND, M.W., MOSMANN, T.R., MOORE, K.W. and RENNICK, D.M., 1991, Interleukin-10: a novel stimulatory factory for mast cells and their progenitors, Journal of Experimental Medicine, 173, 507–10. TOPPING, M.D., VENABLES, K.M., LUCZYNSKA, C.M., HOWE, W. and NEWMAN TAYLOR, A.J., 1986, Specificity of the human IgE response to inhaled acid anhydrides, Journal of Allergy and Clinical Immunology, 77, 834– 42. VANDENPLAS, O., CARTIER, A., LESAGE, J., CLOUTIER, Y., PERREAULT, G., GRAMMER, L.C., SHAUGHNESSY, M.A. and MALO, J-L., 1993, Prepolymers of hexamethylene diisocyanate as a cause of occupational asthma, Journal of Allergy and Clinical Immunology, 91, 850–61. WARDLAW, A.J., 1993, The role of air pollution in asthma, Clinical and Experimental Allergy, 23, 81–96. YING, S., DURHAM, S.R., JACOBSON, M.R., RAK, S., MASUYAMA, K., LOWHAGEN, O., KAY, A.B. and HAMID, Q.A., 1994, T lymphocytes and mast cells express messenger RNA for interleukin-4 in the nasal mucosa in allergen-induced rhinitis, Immunology, 82, 200–6. YOKOTA, T., COFFMAN, R.L., HAGIWARA, H., RENNICK, D.M., TAKEBE, Y., YOKOTA, K., GEMMELL, L., SCHRADER, B., YANG, G., MEYERSON, P., LUH, H., HOY, P., PENE, J., BRIERE, F., SPITS, H., BANCHEREAU, J., DE VRIES, J., LEE, F.D., ARAI, N. and ARAI, K.I., 1987, Isolation and characterization of lymphokine cDNA clones encoding mouse and human IgAenhancing and eosinophil colony-stimulating factor activities. Relationship to interleukin 5. Proceedings of the National Academy of Sciences USA, 84, 7388–92. ZEISS, C.R., KANELLAKES, T.M., BELLONE, J.D., LEVITZ, D., PRUZANSKY, J.J. and PATTERSON, R., 1980, Immunoglobulin E-mediated asthma and hypersensitivity pneumonitis with precipitating anti-hapten antibodies due to diphenylmethane diisocyanate (MDI) exposure, Journal of Allergy and Clinical Immunology, 65, 346–52. ZETTERSTROM, O., OSTERMAN, K., MACHADO, L. and JOHANSSON, S.G., 1981, Another smoking hazard: raised serum IgE concentration and increased risk of occupational allergy, British Medical Journal, 283, 1215–17.

11 Occupational Asthma Induced by Chemical Agents C.A.C.PICKERING Wythenshawe Hospital Manchester

Introduction Occupational asthma may be defined as variable airways narrowing causally related to exposure in the working environment to airborne dust, gases, vapours or fumes (Newman Taylor, 1980). This definition includes, therefore, both immunological and nonimmunological causes of asthma in the workplace. Immunological causes of asthma in general demonstrate a latent period between exposure and the development of symptoms. Once sensitisation has occurred airway responses may be seen at very low levels of exposure. Both high and low molecular weight agents may cause sensitisation. Irritantinduced occupational asthma characteristically follows within 24 h of a usually, single, high level exposure to an irritant substance and has been named reactive airways dysfunction syndrome (Brooks et al., 1985). The number of chemical agents causing occupational asthma is extensive. As new, highly reactive, chemicals are developed these numbers are likely to grow. Low molecular weight chemicals may act as haptens, reacting with body protein to form a complete antigen to which specific antibodies are formed. The incidence of occupational asthma in most countries is not known with any great accuracy, there are considerable variations in reporting systems between countries. Since many individuals with occupational asthma change jobs without a specific diagnosis being established, the published figures of incidence will be significant underestimates of the true incidences. In Japan (Kobayashi et al., 1973), the prevalence of occupational asthma amongst adult male asthmatics is said to be about 15%. In the UK a new reporting system has recently been established— Surveillance of Work-related and Occupational Respiratory Disease Project (SWORD). Newly diagnosed cases of workrelated respiratory disease are reported monthly by consultant chest and occupational physicians. Between 1989 and 1991, 631 cases of chemically induced occupational asthma were reported, of these 53% were associated with exposure to

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isocyanates. A similar system (SHIELD) has been established in the UK in the West Midlands (Gannon et al., 1993). They reported an incidence of 43 new cases per million workers per year. Specific occupational incidences varied from 1833 per million paint sprayers to 8 per million clerks. Again more than half the cases of asthma were attributed to isocyanates. The initial diagnosis of occupational asthma is based on a workers history of respiratory symptoms improving on days away from work and when on holiday. At the onset of occupational asthma this pattern is usually present but continued exposure to the allergen leads to increasing airway reactivity. Their symptoms may then persist over weekends and be triggered by nonspecific factors outside the workplace such as exhaust fumes, aerosol sprays and perfumes. The difficulties of relying on the history alone in the diagnosis of occupational asthma has been well documented (Malo et al., 1991). A series of 162 hospital referrals to two expert physicians were initially categorised, on the basis of their histories, into highly probable, probable, uncertain, unlikely or absent occupational asthma. The diagnosis was then established by bronchial provocation testing and or serial measurements of lung function. The predictive value of a physician’s assessment of occupational asthma being highly probable or probable was only 63%. This improved to 83% in the groups in whom occupational asthma was assessed as being unlikely or absent. The early identification of work-related symptoms and their subsequent investigation in the workplace is important. Only rarely, when very acute episodes of workplace asthma are described, should lung function measurements at work be avoided. While pre- and post-shift measurements of lung function may identify a work-related effect, late asthmatic responses occurring in the evening after leaving work are frequently seen in chemically induced forms of occupational asthma. Serial measurements of lung function made every 2 h from waking to sleeping, both on working days and on days away from work, using a peak flow meter, will identify these late responders. The sensitivity of this type of investigation in establishing a diagnosis of occupational asthma is about 80% (Burge, 1982), this falls to 46% once the worker is started on specific treatment for his asthma, again emphasising the importance of early identification and investigation of work-related respiratory symptoms. Currently more than 140 low molecular weight chemicals have been reported to induce occupational asthma (Butcher and Salvaggio, 1986). The majority of these chemicals induce asthma by mechanisms which have yet to be identified. In a minority of instances specific IgE antibodies to the implicated chemical have been identified. Bronchial challenge tests with chemicals which are non-IgE dependent usually induce either an isolated late asthmatic response or a biphasic or dual asthmatic response. The IgE dependent responses induce immediate or dual asthmatic responses.

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The most common chemical causes of occupational asthma include the iso- cyanates and the acid anhydrides. This chapter will examine these two groups in more detail. Isocyanates The polyisocyanates and their oligomers are the most important cause of chemically induced asthma. These organic compounds are synthesised by the reaction of amines with phosgene. There are a number of related compounds the most important of which are 2,4- and 2,6-toluene diisocyanate (TDI), methylene diphenyldiisocyanate (MDI), hexamethylene diisocyante (HDI), napthalene diisocyanate (NDI), isophorone diisocyanate (IPDI), and polyisocyanates derived from HDI and MDI. The incidence of occupational asthma due to diisocyanates varies widely. It is influenced by the type of compound and its vapour pressure. TDI and HDI are highly volatile at room temperature, whereas MDI has to be heated to above 60°C to volatilise. It is thought that approximately 5% of an exposed working population will develop occupational asthma after exposure to TDI (Diem et al., 1982). Because of the known respiratory problems associated with exposure to isocyanates with high vapour pressure properties, new isocyanate compounds with low vapour pressure properties have been developed particularly for use in the paint spraying industry. Recent studies however continue to demonstrate significant levels of occupational asthma despite the use of recommended respiratory protection (Seguin et al., 1987, Welinder et al., 1988). Bronchial provocation studies with HDI- and MDI-derived polyisocyanates have confirmed their ability to cause occupational asthma. Airborne iso-cyanate prepolymers appear to be able to induce asthma to the same or greater frequency as isocyanate monomers. High exposures to isocyanate vapours, such as occur in a major industrial spillage, cause acute rhinitis, lacrymation, cough and wheezing leading to subsequent sensitisation. In some individuals this type of exposure induces persistent asthma—reactive airways dysfunction syndrome (RADS). Respiratory sensitisation may occur at very low levels of exposure. Pepys et al. (1972) described a boat builder who became sensitised to TDI at exposure levels of between 0.00173 and 0.0018 ppm. Similarly White et al. (1980) reported respiratory symptoms and the development of IgE antibodies to TDI, in machinists manufacturing carseat covers exposed to levels of TDI of between 0.0003 and 0.003 ppm. It is more usual, in the author’s experience, for the sensitised individual to provide a history of short lived peak exposures to isocyanates which have clearly been above the current threshold limit value. These intermittent relatively high level exposures may be important in the sensitisation

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process. Once sensitised, a worker may have his symptoms initiated by very low exposure levels of isocyanates. Diisocyanate asthma is usually but not always associated with the presence of nonspecific bronchial hyperreactivity. The majority of workers who develop occupational asthma remain symptomatic requiring regular treatment permanently after cessation of exposure (Allard et al., 1989). The duration of exposure with symptoms before diagnosis has a major influence on recovery patterns. In a group of 43 isocyanate workers with occupational asthma, those who had fully recovered were exposed with symptoms for 1.6 years, those who had improved, 2.8 years and those who had not improved, 5.4 years (Pisati et al., 1993). The resolution or improvement in occupational asthma takes place over a 2 year period after cessation of exposure, symptoms still present at 2 years should be regarded as permanent. Most epidemiological studies have not identified any specific risk factors including atopic status, smoking or nonspecific bronchial hyperreactivity. The laboratory identification of specific antibodies to diisocyanates has proved of very limited value. Diisocyanate specific IgE is demonstrable in only 10–20% of sensitised individuals and have also been identified in individuals with no history of asthma (Butcher et al., 1983). Similarly specific IgG antibodies to diisocyanates have been described in workers both with and without evidence of disease. At the present time the recommended long-term exposure limit (8 h TWA reference period) for diisocyanates is 0.02 mg m−3 and the short-term exposure limit (10 min reference period) is 0.07 mg m−3 in the UK. There is discussion at the present time as to whether levels should be lower in order to prevent the development of diisocyanate asthma. However since most workers with diisocyanate airways disease describe exposures in excess of the current recommended exposure levels the prevalence of occupational asthma in a workforce without such exposures is not known. There need to be improvements in hygiene control to prevent these peak exposures to isocyanates. Acid anhydrides The acid anhydrides are a group of low molecular weight chemicals used as curing agents in the production of epoxy and alkyd resins and in the production of plasticisers such as dioctyl phthalate. Acid anhydrides exert diverse effects on man both as sensitisers, irritants or both. The most frequently used anhydrides, all of which have been described causing occupational asthma, are phthalic anhydride (PA), trimellitic anhydride (TMA), tetrachlorophthalic anhydride (TCPA) and maleic anhydride (MA). In addition himic anhydride and pyromellitic dianhydride (PMDA) have been described as causing asthma.

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The direct toxicity of anhydrides involves irritation of the mucus membranes and skin which may result in eye lesions, epistaxis, pulmonary congestion, haemoptysis and skin burns (Venables, 1989). Occupational asthma is most frequently reported due to PA, less commonly to TMA, TCPA and MA and finally there are single case reports of asthma due to HA, HHPA and PMDA (Venables, 1989). A second type of response to acid anhydrides has also been described and is termed the ‘late respiratory systemic syndrome’ (LRRS). This is characterised by the development of influenzal type symptoms, fever, generalised acheing and malaise, late in the working shift or in the evening after work. These symptoms may occur in isolation or in association with asthma. It is not clear whether this response is immunologically mediated or a nonspecific response to high levels of anhydride exposure. Lastly, exposure to TMA, probably at high exposure levels, has been described as causing severe pulmonary haemorrhage requiring both blood transfusion and mechanical ventilation (Rivera et al., 1989). Serum IgE and IgG antibodies to acid anhydrides have been identified. IgE antibodies appear to be more specifically associated with occupational asthma. Howe et al. (1983) reported seven cases of TCPA asthma all of whom had IgE antibody to TCPA, compared with 8% of 300 exposed workers without TCPA asthma; 29% of this exposed nonasthmatic population had IgG antibodies to TCPA. The exposure levels of acid anhydrides that initiate sensitisation are poorly understood. TMA at levels of 1.7–4.7 mg m−3 (Zeiss et al., 1977) and 0.007–2.1 mg m−3 (Bernstein et al., 1983; McGrath et al., 1984) have been described causing occupational asthma. PA at 0.03–15 mg m−3 has also been reported as causing asthma (Wernfors et al., 1986). As in other forms of occupational asthma, the early identification of cases of acid anhydride induced asthma and their removal from exposure is of prime importance. Reactive airways dysfunction syndrome Reactive airways dysfunction syndrome (RADS) or irritant-induced asthma was first described in 1985 (Brooks et al., 1985). The criteria used in diagnosis include a high level exposure to an irritant fume, vapour or smoke, the development of respiratory symptoms within minutes or hours of exposure, in an individual with no previous history of respiratory symptoms, with persistence of symptoms and physiological abnormalities for more than 1 year. A variety of different chemical exposures have been described inducing this syndrome including: chlorine (Moore and Sherman, 1991), glacial acetic acid (Kern, 1991), hydrochloric acid (Promisloff et al., 1990) and miscellaneous chemical exposures (Brooks et al., 1985). A comparison between cases of occupational asthma and RADS (Gautrin et al., 1994)

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suggest that cases with RADS are left with less airway reversibility than occupational asthmatics. This would be consistent with the pathological findings (Boutet et al., 1993) in RADS, with more severe basement membrane thickening and bronchial wall fibrosis than is present in occupational asthma. The development of occupational asthma in any individual has potentially serious consequences both in terms of persisting disability, possible unemployment and loss of income (Gannon et al., 1993). It is incumbent on management to ensure safe working conditions with adequate control and regular monitoring of atmospheric levels of chemical agents. References ALLARD, C., CARTIER, A., GHEZZO, H. and MALO, J-L., 1989, Occupational asthma due to various agents. Absence of clinical and functional improvement at an interval of four or more years after cessation of exposure, Chest, 96, 1046–9. BERNSTEIN, D.I., ROACH, D.E., MCGRATH, K.G., LARSEN, R.S., ZEISS, C. R. and PATTERSON, R., 1983, The relationship of airborne trimellitic anhydrideinduced symptoms and immune responses, J. Allergy Clin. Immunol., 72, 709–13. BOUTET, M., BOULET, L.-P., MALO, J.L., CARTIER, A., CÔTÉ, J., LEBLANC, C., MILOT, J. and LAVIOLETTE, M., 1993, Morphological evidence of modified contractile properties of airways in occupational asthma and reactive airways dys-function syndrome, Am. Rev. Respir. Dis., 147, A113. BROOKS, S.M., WEISS, M.A. and BERNSTEIN, I.L., 1985, Reactive airways dysfunction syndrome. Case reports of persistent airways hyperreactivity following high-level irritant exposures, J. Occup. Med., 27, 473–6. BURGE, P.S., 1982, Single and serial measurements of lung function in the diagnosis of occupational asthma, Eur. J. Resp. Dis., 63 (suppl. 123), 47–9. BUTCHER, B.T. and SALVAGGIO, J.E., 1986, Continuing medical education— occupational asthma, J. Allergy Clin. Immunol. 78, 547–9. BUTCHER, B.T., O’NEIL, C.E., REED, M.A. and SALVAGGIO, J.E., 1983, Radioallergosorbent testing with p-tolyl monoisocyanate in toluene diisocyanate workers, Clin. Allergy., 13, 31–4. DIEM, J.E., JONES, R.N., HENDRICK, D.J., GLINDMEYER, H.W., DHARMARAJAN, V., BUTCHER, B.T., SALVAGGIO.J.E., and WEILL, H., 1982, Five year longitudinal study of workers employed in a new toluene diisocyanate manufacturing plant, Am. Rev. Respir. Dis., 126, 420–8. GANNON, P.F.G. and BURGE, P.S., 1993, The SHIELD scheme in the West Midlands region, United Kingdom, Brit. J. Ind. Med., 50, 791–6. GANNON, P.F.G., WEIR, D.C., ROBERTSON, A.S. and BURGE, P.S., 1993, Health, employment, and flnancial outcomes in workers with occupational asthma, Brit. J. Ind. Med., 50, 491–6.

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GAUTRIN, D., BOULET, L.-P., BOUTET, M., DUGAS, M., BHÉRER, L., L’ARCHEVÊQUE, J., LAVIOLETTE, M., CÔTÉ, J. and MALO, J.-L., 1994, Is reactive airways dysfunction syndrome a variant of occupational asthma? J. Allergy Clin. Immunol, 93, 12–22. HOWE, W., VENABLES, K.M., TOPPING, M.D., DALLY M.B., HAWKINS, R., LAW, J.S. and NEWMAN TAYLOR, A.J., 1983, Tetrachlorophthalic anyhdride asthma: evidence for specific IgE antibody, J. Allergy Clin. Immunol., 71, 5–11. KERN, D.G., 1991. Outbreak of the reactive airways dysfunction syndrome after a spill of glacial acetic acid, Am. Rev. Respir. Dis. 144, 1058–64. KOBAYASHI, S.Y., YAMAMORA, Y., FRICK, O.L., HORIUCHI, S. KISHIMOTO, T. and MIYAMOTO, T., 1973, Occupational asthma due to the inhalation of pharmaceutical dusts and other chemical agents with some reference to other occupational asthma in Japan, Proc. VIII Int. Congr. Allergology, Tokyo, October 1973, pp. 124–32. Amsterdam: Excerpta Medica. MALO, J.L., GHEZZO, H., L’ARCHEVÊQUE, LAGIER, F. and CARTIER, A., 1991, Is the clinical history a satisfactory means for diagnosing occupational asthma? Am. Rev. Respir. Dis., 143 528–32. MCGRATH, K.G., ROACH, D., ZEISS, C.R. and PATTERSON, R., 1984, Four year evaluation of workers exposed to trimellitic anhydride: a brief report, J. Occup. Med., 26, 671–5. MOORE, B.B. and SHERMAN, M., 1991, Chronic reactive airway disease following acute chlorine gas exposure in an asymptomatic atopic patient, Chest, 100, 855–6. NEWMAN TAYLOR, A.J., 1980, Occupational asthma, Thorax, 35, 241–5. PEPYS, J., PICKERING, C.A .C., BRESLIN, A.B.X. and TERRY D.J., 1972, Asthma due to inhaled chemical agents—tolylene diisocyanate, Clin. Allergy, 2. 225–36. PISATI, G., BARUFFINI, A. and ZEDDA, S., 1993, Toluene diisocyanate induced asthma: outcome according to persistence or cessation of exposure, Brit. J. Ind. Med., 50, 60–4. PROMISLOFF, R.A., PHAN, A., LENCHNER, G.S. and CICHELLI, A.V., 1990, Reactive airway dysfunction syndrome in three police officers following a roadside chemical spill, Chest, 98, 928–9. RIVERA, M., NICOTRA, M.B., BYRON, G.E., PATTERSON, R., YAWN, D.H., FRANCO, M., ZEISS, C.R. and GREENBERG, S.D., 1981, Trimellitic anhydride toxicity: a cause of acute multisystem failure, Arch. Intern. Med., 141, 1071– 4. SEGUIN, P., ALLARD, A., CARTIER, A. and MALO, J.-L., 1987, Prevalence of occupational asthma in spray painters exposed to several types of isocyanates, including polymethylene polyphenylisocyanate, J. Occup. Med., 29, 340–4. VENABLES, K.M., 1989, Low molecular weight chemicals, hypersensitivity, and direct toxicity: the acid anhydrides, Brit. J. Ind. Med., 46, 222–32. WELINDER, H., NIELSEN, J., BENSRYD, I. and SKERFVING, S., 1988, IgG antibodies against polyisocyanates in car painters, Clin. Allergy, 18, 85–93.

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WERNFORS, M., NIELSEN, J. and SKERFVING, S., 1986, Phthalic anhydrideinduced occupational asthma, Int. Arch. Allergy Appl. Immunol. 79, 77–82. WHITE, W.G., SUGDEN, E., MORRIS, M.J. and ZAPATA, E, 1980, Isocyanateinduced asthma in a car factory, Lancet, 1, 756–60. ZEISS, C.R., PATTERSON, R., PRUZANSKY, J.J., MILLER, M.M., ROSENBERG, M. and LEVITZ, D., 1977, Trimellitic anhydride-induced airway syndromes: clinical and immunological studies, J. Allergy Clin. Immunol., 60, 96–103.

PART FOUR Biomarkers and risk assessment of industrial chemicals

12 Biomarkers and Risk Assessment KARI HEMMINKI Karolinska Institute, Huddinge

Introduction Many chemical carcinogens cause covalent DNA-binding products, adducts, which may induce mutations or other types of DNA damage in important growth-controlling genes or loci resulting in aberrant cellular growth and cancer (Harris, 1991; IARC 1992; Hemminki, 1993). Human exposure to compounds such as polycyclic aromatic hydrocarbons (PAH) can be determined, for example, by ambient air, biological or DNA adduct monitoring. The usefulness of a method for the determination of DNA adducts in human biomonitoring requires high sensitivity because the levels of adducts are low. Here the primary focus is on the assessment of exposure using the above indicators in industries where high exposure to PAHs occur, such as iron founding, coke production, aluminium production, garage work and engine overhauling with exposure to used lubricating oils. Biomonitoring of PAH exposure Literature on the application of DNA adduct studies in humans is extensive (Beach and Gupta, 1992; IARC, 1993, 1994; Hemminki et al., 1993a; Hemminki, 1994). A large majority of the 32P-postlabelling studies on human samples focus on tobacco smoking, occupational exposures and cancer chemotherapy patients. Most occupational exposures studied relate to complex mixtures, including polycyclic aromatic hydrocarbons (PAHs). In exposure to complex mixtures multiple radioactive spots (called diagonal radioactive zones, DRZ) are detected. The adduct spots cannot be definitively identified nor quantitated. As it has turned out that for many adducts labelling is not completed, even among structural analogues such as PAHs, the adduct levels measured are likely to be underestimates (Segerbäck and Vodicka, 1993).

Notes: a Total white blood cells. b Lymphocytes. c Not given. d No data.

Table 12.1 Exposure and aromatic adducts in occupational populations, presented in simplified tabulated form

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The types of occupational groups studied by postlabelling include foundry, coke oven and aluminium workers, roofers, garage and terminal workers, car mechanics and chimney sweeps. All these groups have had an

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increased risk of lung cancer. As, however, epidemiological studies relate to exposure a few decades earlier the risks of present exposures can only be predicted. The levels of aromatic adducts are elevated in white blood cells or lymphocytes in many of these groups. The reported total aromatic adduct levels usually range between 1 and 10 adducts per 108 nucleotides. Long-lived lymphocytes tend to have higher levels of adducts than shortlived granulocytes (Savela and Hemminki, 1991; Grzybowska et al., 1993). As a rule of thumb, it can be assumed that in a steady-state (i.e. long term exposure) lymphocytes contribute to the level of adducts overwhelmingly. Because they represent about 25 percent of the DNA in blood, the relationship between total white blood cell (WBC) and lymphocyte DNA adducts should be about 1:4, granulocytes only contributing to the amount of DNA denominator. Yet one has to be cautious in the comparison of results between various assays even within a laboratory as the results may ‘drift’ with time. The levels of white blood cell/lymphocyte aromatic adducts from workers in several industries, as measured by postlabelling, and as compared to ambient air concentrations of benzo (a) pyrene (BP) and 1hydroxypyrene levels are presented in Table 12.1. A boxplot presentation of the adduct levels of bus maintenance and truck terminal workers is shown in Figure 12.1 (Hemminki et al., 1994). The differences that were statistically significant from the controls were, in addition to the groups of maintenance and terminal workers, garage workers and diesel forklift drivers. There does not seem to be a direct relationship between exposure and adduct levels. Electrode, coke and aluminium workers, exposed up to several 100 ng m−3 concentrations of BP, do not differ from the control more than foundry workers, exposed to less than 1/10 of the cited levels. The apparently higher level of adducts in the aluminium and electrode former workers (and controls) as compared to the other measurements, is due a method applied earlier with higher amounts of radioactive ATP. The later assays were carried out in small volumes but high concentrations of ATP (Hemminki et al., 1993b; Szyfter et al., 1994). An increased level of lymphocyte adducts has also been found in garage and truck terminal workers, with estimated exposures of about 10 ng m−3 (Hemminki et al., 1994). This would imply that the detection limit of the postlabelling method in humans exposed to PAHs lies somewhere between 1 and 10 ng m−3 BP. Whether diesel exhaust is a particularly potent inducer of adducts remains to be demonstrated. The differences between the exposed and the controls are statistically significant among foundry workers, all bus maintenance personnel and garage workers as a subgroup, all truck terminal workers and the diesel forklift drivers in particular. Coke workers differed significantly from the local controls in summer when

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Figure 12.1 A boxplot of the white blood cell DNA adduct levels (per 108 nucleotides) among bus maintenance and truck terminal workers and controls (Hemminki et al., 1994).

environmental pollution was low and the adduct levels in the controls were about 1/10 of their level in the winter (Grzybowska et al., 1993). Adducts and other endpoints It has become customary to include many types of endpoints to biomonitoring studies. The foundry study cited in Table 12.1 belongs to the most versatile of them. Exposure is measured by ambient air and 1hydroxypyrene monitoring (Santella et al., 1993). DNA adducts are assayed for by postlabelling and immunoassay. Plasma albumin PAH adducts are measured. Hypoxanthin guanine phosphoribosyl transferase (HPRT) and glycophorin A mutations are assayed for in lymphocytes and erythrocytes, respectively (Perera et al., 1993, 1994). Single-stand breaks in DNA and three types of cytogenetic parameters, chromosomal aberrations, sister chromatid exchanges and micronuclei, are analysed, in addition to genotyping of drug metabolising enzyme genes. Sampling of workers was repeated in four consecutive years, each at the same time of the year. As the last sampling was in the end of 1993, it will take some time before the complete data set will be available for analysis. In some published work from this data set an increase in DNA adducts and mutation frequency in the HPRT and glycophorin A genes was reported (Figure 12.2). Yet unreported results appear to show an increase

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Figure 12.2 Total white blood cell DNA adducts, determined by immunoassay, HPRT and glycophorin A mutations in foundry workers, exposed to various levels of BP (Perera et al., 1993).

in singlestrand breaks while none of the cytogenetic parameters are elevated in the foundry workers. Adducts and metabolic genotypes The modulation of environmental carcinogenesis by host polymorphism in genes for xenobiotics metabolising enzymes is currently under extensive investigation. It was initially sparked by findings linking certain phenotypes of drug metabolism to cancer risk (Seidegård et al., 1986; Nebert, 1991). The enzymes of interest in the context of exposure to PAHs include cytochrome P450 CYP1A1 and glutathione transferase GST, involved in the activation and inactivation, respectively, of PAHs. By restriction enzyme mapping two allelic forms, ml and m2, and two other closely linked codons for isoleucine (Ile, linked to ml) and valine (Val, linked to m2) can be defined, where m2 and valine represent the rare mutant genotypes, associated with the inducibility of the enzyme activity (Hayashi et al., 1991). Polymorphism in GST1 involves the presence or the absence of the gene (Nakachi et al., 1992). The null genotype lacks the enzyme completely. Among chimney sweeps there was an association of the rare, inducible CYP1A1 genotype ml/m2 with low adduct levels in white blood cell DNA (Ichiba et al., 1994). In the same study an increased level of adducts was noted in the GST1-individuals. The level of DNA adducts appeared to be related to both the GST and CYP1A1 genotype (Figure 12.3). Analysis of micronuclei in chimney sweeps resulted in no differences between individuals of either CYP1A1 ml/m2, m2/m2 or Ile/Val genotypes nor of GST1 + or − genotypes (Carstensen et al., 1993). There was however a

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Figure 12.3 Total white blood cell DNA adducts, measured by postlabelling, according to CYP1A1 and GST1 genotype (Ichiba et al., 1994). Controls, Sweeps.

correlation between white blood cell DNA adducts and micronuclei and it was stronger among the GST-individuals (Ichiba et al., 1994). How important is the role of metabolic phenotype or genotype as a predictor of cancer risk remains to be established. However it would seem prudent to assume some role as long as there is significant exposure to a carcinogen, metabolism of which is regulated by polymorphic genes. It would be important to note that the question can only be addressed if both of these conditions are met. In much of the published literature there are uncertainties regarding the active agents and their metabolic routes in the tissues studied. Adjustment for a metabolic phenotype or genotype, when justified, may increase the precision in the measurement. Risk assessment Monitoring of DNA adducts in occupational setting has mainly been applied to workers exposed to PAHs. In the case of 32P-postlabelling increases in the level of adducts has been noted at exposures around 10 ng BP m−3 or slightly below. This is close to the detection limit that can conveniently be attained with personal monitoring or by measuring urinary 1-hydroxypyrene. As the adduct measurements also reflect some aspects of metabolism and DNA repair, they extend the scope of exposure

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measurements to host factors that may underly individual susceptibility to cancer. It has become increasingly common to try and incorporate other endpoints to DNA adduct studies. These include metabolic parameters, discussed above, protein adducts, cytogenetic parameters and point mutations. Examples include ethylene oxide exposed workers (Tates et al., 1991) and foundry workers (Perera et al., 1993, 1994; Santella et al., 1993). In both studies several parameters were elevated. The study on chimney sweeps illustrated how the intermediary endpoint may increase precision in the measurements (cf. Figure 12.3). The initial study showed no correlation between sweeping and micronuclei even though an adjustment was made for CYP1A1 and GST genotypes (Carstensen et al., 1993). There was a moderate correlation between sweeping and white blood cell DNA adducts, and adducts and micronuclei. Both of these correlations were strengthened once GST genotype was considered (Ichiba et al., 1994). Increasing circumstantial evidence associates DNA adducts within groups to an increased risk of cancer (IARC, 1992; Hemminki, 1993). Many of the adduct studies have been carried out in occupational groups which have been at a risk of cancer based on epidemiological results. These studies may be old and relate to exposures decades ago. Even new epidemiological publications on cancer cannot accurately address exposures after about 1970. Simultaneously there have been large changes in technology and industrial hygiene, undermining the quantitative and sometimes even the qualitative findings of the old epidemiological studies. This is one justification for the biomonitoring studies. Another justification is on exposures where epidemiological studies have not been conducted or have provided inadequate results, in spite of suspicions raised by short-term or animal experiments. The International Agency for Research on Cancer has pointed out this as one of the criteria to be used in the evaluation of carcinogenicity of chemicals (IARC, 1992). Styrene belongs to this group of industrial exposures, where epidemiological findings are equivocal but adduct data are available on workers (Vodicka et al., 1993). Acknowledgements The research was supported by the Swedish Medical Research Council and Work Environment Fund.

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References BEACH, A. and GUPTA, R., 1992, Human biomonitoring and the 32P-postlabeling assay, Carcinogenesis, 13, 1053–74. CARSTENSEN, U., ALEXANDRIE, A.-K., HÖGSTEDT,B., RANNUG, A., BRATT, I. and HAGMAR, L., 1993, B- and T-lymphocyte micronuclei in chimney sweeps with respect to genetic polymorphism for CYP1A1 and GST1 (class Mu), Mutat. Res., 289, 187–95. GRZYBOWSKA, E., HEMMINKI, K., SZELINGA, J. and CHORAZY, M., 1993, Sea-sonal variation of aromatic DNA adducts in human lymphocytes and granulocytes, Carcinogenesis, 14, 2523–6. HARRIS, C., 1991, Chemical and physical carcinogenesis: advances and perspectives for the 1990s, Cancer Res., 51, 5023–44. HAYASHI, S., WATANABE, J., NAKACHI, K. and KAWAJIRI, K., 1991, Genetic linkage of lung cancer-associated Msp I polymorphisms with amino acid replacement in the heme binding region of the human cytochrome P4501A1 gene, J. Biochem., 110, 407–11. HEMMINKI, K., 1993, DNA adducts, mutations and cancer, Carcinogenesis, 14, 2007– 12. HEMMINKI, K., 1995, DNA adducts in biomonitoring, J. Occup. Environ. Med., 37, 44–51. HEMMINKI, K., AUTRUP, H. and HAUGEN, A., 1993a, Environmental carcinogens: Assessment of Exposure and Effect, pp. 89–102, Heidelberg: Springer Verlag. HEMMINKI, K., FÖRSTI, A., LÖFGREN, M., SÄGERBÄCK, D., VACA, C. and VODICKA, P., 1993b, Testing of quantitative parameters in the 32Ppostlabelling method, in Phillips, D.H., Castegnaro, M. and Bartsch, H. (Eds), Postlabelling Methods for Detection of DNA Adducts, IARC Sci. Publ., No. 124, pp, 51–63, Lyon: IARC. HEMMINKI, K., SÖDERLING, J., ERICSON, P., NORBECK, H.E. and SEGERBÄCK, D., 1994, DNA adducts among personnel servicing and loading diesel vehicles, Carcinogenesis, 15, 767–9. IARC, 1992, Mechanisms of Carcinogenesis in Risk Identification, IARC Sci. Publ. No. 116, Lyon: IARC. IARC, 1993, Postlabelling Methods for Detection of DNA Adducts, IARC Sci. Publ. No. 124, Lyon: IARC. IARC, 1994, DNA Adducts: Identification and Biological Significance, IARC Sci. Publ. No. 125, Lyon: IARC. ICHIBA, M., HAGMAR, L., RANNUG, A.O., HÖGSTEDT, B., ALEXANDRIE, A.K. and HEMMINKI, K., 1994, Aromatic DNA adducts, micronuclei and genetic polymorphism for CYP1A1 and GST1 in chimney sweeps, Carcinogenesis, 15, 1347–52. NAKACHI, K., IMAI, K., HAYASHI, S.I., WATANABE, J., KAWAJIRI, K., HAYASHI, S.-I., WATANABE, J. and KAWAJIRI, K., 1992, High susceptibility to lung cancer analyzed in terms of combined genotypes of P450IA1 and mu-class glutathione S-transferase genes, Jpn J. Cancer Res., 83, 866–70.

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NEBERT, D.W., 1991, Role of genetics and drug metabolism in human cancer risk, Mutat. Res., 247, 267–81. ØVREBø, S., HAUGEN, A., FJELDSTAD, P.E., HEMMINKI, K. and SZYFTER, K., 1994, Biological monitoring of exposure to PAH in an electrode paste plant, J. Occup. Med., 36, 303–10. ØVREBø, S., HAUGEN, A., HEMMINKI, K., SZYFTER, K., DRABLÖS, P.A. and SKOGLAND, M., 1995, Studies of biomarkers in aluminium workers occupationally exposed to polycyclic aromatic hydrocarbons, Cancer Detection Prev., 19, 258. PERERA, F.P., TANG, D.L., O’NEILL, J.P., BIGBEE, W.L., ALBERTINI, R.J., SANTELLA, R., OTTMAN, R., TSAI, W.Y., DICKEY, C., MOONEY, L.A., SAVELA, K. and HEMMINKI, K., 1993, HPRT and glycophorin A mutations in foundry worker: relationship to PAH exposure and PAH-DNA adducts, Carcinogenesis, 14, 969–73. PERERA, F.P., DICKEY, C., SANTELLA, R., O’NEILL, J.P., ALBERTINI, R.J., OTTMAN, R., TSAI, W.Y., MOONEY, L.A., SAVELA, K. and HEMMINKI, K., 1994, Carcinogen-DNA adducts and gene mutations in foundry workers with changing exposure to PAH, Carcinogenesis, 15, 2905–10. SANTELLA, R.M., HEMMINKI, K., TANG, D.-L., PAIK, M., OTTMAN, R., YOUNG, T.L., SAVELA, K., VODICKOVA, L., DICKEY, C., WHYATT, R. and PERERA, P.P., 1993, Polycyclic aromatic hydrocarbon-DNA adducts in white blood cells and urinary 1-hydroxypyrene in foundry workers, Cancer Epidemiol. Biomarkers Prevent., 2, 59–62. SAVELA, K. and HEMMINKI, K., 1991, DNA adducts in lymphocytes and granulocytes of smokers and non-smokers detected by the 32P-postlabelling assay, Carcinogenesis, 12, 503–8. SEGERBÄCK, D. and VODICKA, P., 1993, Recoveries of DNA adducts of polycyclic aromatic hydrocarbons in the 32P-postlabelling assay, Carcinogenesis, 14, 2463–9. SEIDEGǺRD, J., PERO, R.W., MILLER, D.G. and BEATTIE, E.J., 1986, A glutathione transferase in human leukocytes as a marker for the susceptibility to lung cancer, Carcinogenesis, 7, 751–3. SZYFTER, K., KRUGER, J., ERICSON, P., VACA, C., FÖRSTI, A. and HEMMINKI, K., 1994, 32P-postlabelling analysis of DNA adducts in humans: adducts distribution and method improvement, Mut. Res., 313, 269–76. TATES, A.D., GRUMMT, T., TÖRNQVIST, M., FARMER, P.B., VAN DAM, F.J., VAN MOSSEL, H., SCHOEMAKER, H.M., OSTERMAN-GOLKAR, S., UEBEL, C., TANG, Y.S., ZWINDERMAN, A.H., NATARAJAN, A.T. and EHRENBERG, L., 1991, Biological and chemical monitoring of occupational exposure to ethylene oxide, Mut. Res., 250, 483–97. VODICKA, P., VODICKOVA, L. and HEMMINKI, K., 1993, 32P-postlabelling of DNA adduct of styrene-exposed lamination workers, Carcinogenesis, 14, 2007–12.

13 Extrapolation of Toxicity Data and Assessment of Risk NORBERT FEDTKE Hüls AG, Marl

Introduction Risk assessment provides a link between scientific research and risk management, or in other words, it is ‘a method for reaching public policy decisions’ (Silbergeld, 1993). Risk assessment includes the key elements hazard identification, dose-response assessment, and exposure assessment. These elements are integrated in a risk characterization step to predict adverse effects that may occur in a given population in a particular exposure situation, often based on the quantification of the likelihood of this occurrence. Risk management determines whether the particular exposure situation presents an acceptable or unacceptable risk and whether it is necessary to reduce the risk by reducing the exposure. Whereas risk management has to account for public health, socio-economical factors, technical feasibility, social perceptions, governmental policy and political consequences, risk assessment should be based on scientific principles. Since for the majority of industrial chemicals no or only limited human data exist, the question of how to extrapolate the data obtained from laboratory studies in experimental animals in order to predict the effects in humans has become one important aspect in risk assessment. The final goal is either to determine a level of exposure at which there is no reasonable doubt that an adverse effect will not occur in man or to define the risk associated with this exposure level. The use of mechanistic information to provide linkages between exposure, dose to tissue, and biological responses may assist in some of the steps necessary in the process of species extrapolation. Especially the use of physiologically based pharmacokinetic models (PBPK) for some aspects of risk assessment has been promoted to reduce the uncertainty associated with the current default methods. PBPK modeling is explained in general terms and a recently developed PBPK model for 2-butoxyethanol is provided as an example to illustrate the use of kinetic and mechanistic data in risk assessment.

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Legislative background in the European Union Although the process of risk assessment is not new for the individual member states of the European Union (EU) and representatives of industry and government have been assessing risk for human health and the environment for decades, EU legislation for existing1 and new2 substances did not formally require a systematic risk assessment up to 1992. The situation has changed with the Seventh Amendment of the Directive on the Classification, Packaging and Labelling of Dangerous Substances (EEC, 1992) and the existing substances regulation (EEC, 1993a), which address risk assessment of new and existing substances, respectively. For new chemicals, the general principles of risk assessment are defined in Commission Directive 93/67/EEC (EEC, 1993b). In addition, the Directorate General XI of the European Commission has issued a series of draft guidance documents for use by the competent authorities appointed by the member states. These documents provide the technical details for the risk assessment of new substances mainly by defining the testing strategies for individual toxic endpoints. For existing chemicals, a guidance document has been drafted. However, these technical guidance documents provide only little information on how to extrapolate laboratory data to humans. It has to be assumed that the extrapolation principles in the EU member states will be based on historical approaches used by authorities in other countries. Approaches to risk assessment The final goal of species extrapolation is to define a dose or dose rate which produces no adverse effects in humans. The estimation of a human no-effectlevel may include: – determination of the appropriate animal species for extrapolation to man, – determination of the most critical effect(s) and the target organ(s), – determination of the no-observed-adverse-effect level(s) (NOAEL) of this effect(s), often followed by – extrapolation of the NOAEL(s) from subacute or subchronic to chronic exposure (time extrapolation), – extrapolation of effects observed in a high-dose region of a dose response curve to a low-dose region,

1

Listed in the European Inventory of Existing Commercial Substances (EINECS). 2 Not listed in EINECS.

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– extrapolation of effects from one route of exposure to another route, and – extrapolation of effects observed in a rather homogeneous animal population to a heterogeneous human population (interspecies extrapolation), which also has to take into account the existence of subgroups regarded as more sensitive as the rest of the population (intraspecies extrapolation). Essential for all procedures used in health risk assessment is the determination of the so-called critical effect. The critical effect may be defined as the adverse effect judged to be most appropriate as the basis for the risk assessment. Hence, the first step is the review of all available data on a chemical and the assessment of the adequacy of the database for the determination of the critical effect. On the basis of the critical effect, toxicants may be divided into two classes characterized by: – a threshold of response, i.e. the adverse effect on health is not expressed until the chemical, or the ultimately toxic metabolite, reaches a threshold dose or dose rate in the target tissue, or – no threshold of response, i.e. there is no threshold exposure level below which effects will not be expressed. This implies that there is some risk at any level of exposure. Examples are genotoxic carcinogens or germ cell mutagens. Based on these classes two general approaches to health risk assessment have been used. The first approach involves the use of ‘safety factors’ applied to the NOAEL or the lowest-observed-adverse-effect level (LOAEL) of a threshold effect determined in experimental animals (safety factors are recently referred to as ‘uncertainty’ or ‘assessment’ factors). The magnitude of the uncertainty factors varies between the regulatory bodies that are concerned with risk assessment, but usually they take into account the interspecies extrapolation (default factor 10) and intraspecies extrapolation (default factor 10). The magnitude of the default factors appears to be based more on the conventional use of the decimal system than on scientific reasons and have been proposed first by Lehman and Fitzhugh (1954) for the derivation of acceptable daily intakes (ADIs) for food additives. Additional uncertainty factors may be used for extrapolation to chronic exposure from subacute or subchronic exposure, adequacy of the database, extrapolation of a LOAEL to a NOAEL and severity of effects. The resulting overall uncertainty factor often reaches values of 1000 or higher, which is an indication of the imprecision of the derived tolerable intake. Refined extrapolation procedures using subdivisions of the default

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factors or different default factors have recently been published (Lewis et al., 1990; Renwick, 1991, 1993). The second approach (for non-threshold effects) also relies mainly on default assumptions for dose-response extrapolation and cross-species extrapolation. Especially cancer risk assessment has been the subject of much debate and there are a number of extrapolation methods reviewed recently by Park and Hawkins (1993) and Hallenbeck (1993). The default methodology in the . has been summarized by Frederick (1993). In principle, the risk assessment is based on a chronic rodent bioassay conducted at or near the maximum tolerated dose (MTD). The lifetime constant dose rates and the tumour incidence data for the individual dose groups are used to determine the dose response by fitting the data with a computer program. The linearized multistage cancer model (LMS) is often used to perform this step. The LMS model extrapolates the rodent tumor data observed at the MTD to a dose with a predefined risk and the 95 per cent upper bound on the dose-response curve is calculated. The interspecies extrapolation to humans is performed by a correction factor based on body weight or body surface. Subsequently, the dose is determined that corresponds to a maximum allowable calculated upper bound on risk. The resulting number does not describe the actual human risk under low-level environmental exposure, but provides an upper bound to human risk that is assumed not to be exceeded. The actual risk may be in the range between 0 and the upper bound. In the process described, the dose is defined as administered dose or inhaled concentration. As a result, the lowdose extrapolation does not take into account non-linearities in tissue dosimetry and response. In addition, the interspecies extrapolation is performed using a default approach that does not account for mechanistic species differences. Use of PBPK models in risk assessment General description Physiologically based pharmacokinetic (PBPK) models have been used increasingly over the past decade to improve several aspects of the assessment of risk associated with human exposure to chemicals. Examples are PBPK models for styrene (Ramsey and Andersen, 1984; Csanády et al., 1994), dichloromethane (Andersen et al., 1987), 1,4-dioxane (Reitz et al., 1990a), chloroform (Reitz et al., 1990b), ethyl acrylate (Frederick et al., 1992), methanol (Horton et al., 1992) and 1,3-butadiene (Johanson and Filser, 1993). Recent reviews of the use of PBPK models in risk assessment have been published by several authors (Frederick, 1993; Travis, 1993; Wilson and Cox, 1993; Andersen and Krishnan, 1994).

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PBPK models are based on the blood and tissue solubility of chemicals, their metabolism in various tissues and the physiology of the organism, thus incorporating the specific physiological description of animal species as well as specific physico-chemical descriptions of agents. Uptake, distribution, metabolism and excretion are described in physiologically realistic compartments (tissue groups) using computer simulation. The compartments are linked in parallel, represent the actual mammalian architecture, and include tissues such as lung and arterial blood, fatty tissue, poorly perfused tissues (muscles, skin), richly perfused tissues (brain, kidneys, heart, endocrine gland, gastro-intestinal tract), liver as the main metabolizing tissue, and mixed venous blood. The compartments are connected by arterial and venous blood flow and are characterized by a set of mass balance differential equations. The rate constants that describe the flow of material between the tissue groups and the rate of change in the chemical concentration of each compartment are proportional to blood flow, tissue solubility and compartment volumes. The basic mathematical description of a PBPK model for a volatile compound has been provided by Ramsey and Andersen (1984) and additional details may be found in appendices of manuscripts dealing with the development of PBPK models. Estimation of the constants used in PBPK models may be based on the literature in the case of physiological parameters such as ventilation rates, cardiac output, blood flow to tissues and tissue volumes. The EPA has compiled reference values for these parameters and their scaling (Arms and Travis, 1988). Chemical specific parameters such as blood and tissue solubilities may be determined from in vitro preparation (Sato and Nakajima, 1979; Gargas et al., 1989). The biochemical constants for metabolism may be derived from in vitro studies (Reitz et al. 1988; Carfagna and Kedderis, 1992; Johanson and Filser, 1993), in vivo toxicokinetic studies (Potter and Tran, 1993; Frederick et al., 1992) or in the case of volatile substances from gas uptake studies (Gargas et al., 1986, 1990; Filser, 1992). Since the tissue groups have a defined biological meaning, scaling of the associated parameters between species is possible since many of the parameters used are correlated to body weight. Cardiac output, alveolar ventilation rate and Vmax are scaled by the 3/4 power of body weight whereas Km is assumed to be constant across species. However, the substitution of the physiological parameters with the appropriate values characteristic for the species of interest is preferred. The development of PBPK models is an iterative process involving comparison of the model simulations with experimental data and refinement of the estimates when the model fails to accurately predict the kinetic behaviour. Different exposure scenarios can be used to predict the concentrations of the parent chemical or its metabolites in the blood or the tissues, which are the target of toxic effects. The level of glutathione

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depletion in hepatic and extrahepatic tissues (D’Souza et al., 1988; Frederick et al., 1992; Krishnan et al., 1992), kinetic interactions of parent compounds in mixed exposures (Tardif et al., 1993) or the amount of adducts formed by macromolecular binding (Krishnan et al., 1992) are predictions that may also be generated by PBPK modeling. As a result of the simulations, quantitative information on the internal dose of a chemical or its metabolites in the target tissue is obtained and can replace the administered dose conventionally used in risk assessment. After validation of the PBPK models in experimental animals, human PBPK models can be developed either by allometric scaling of the physiological and biochemical parameters or preferably using the actual human parameters. Following the prediction of the target tissue dosimetry in humans, the appropriate dose surrogates are related to the effect of interest and quantitative species differences are determined. This information provides the possibility to base the species extrapolation on scientific data instead of on arbitrarily assigned default factors and as a consequence the uncertainty of the extrapolation procedures applied in conventional risk assessment may be reduced. Description and use of the PBPK model for 2butoxyethanol 2-Butoxyethanol (BE) is a widely produced glycol ether used as a key ingredient in water- or solvent-based coatings, industrial and consumer cleaning products, and as solvent in a variety of products. Haemolysis was identified as most sensitive indicator of BE-induced toxicity in several species of laboratory animals and has received the most attention as a critical effect for human risk assessment (ECETOC, 1985, 1994). The experimentally determined subchronic NOAEL for the rat is 25 ppm. The major metabolite of BE is 2-butoxyacetic acid (BAA) which has been identified as the metabolite responsible for the haemolysis of red blood cells in in vitro and in vivo studies (Bartnik et al., 1987; Ghanayem et al., 1987; Ghanayem, 1989). Changes in the deformability of rat erythrocytes appear to precede haemolysis upon treatment with BAA. Treatment of human erythrocytes with BAA did not induce changes in deformability (Udden and Patton, 1994; Udden, 1994). The observed species differences may be due to differences in the lipid composition of erythrocyte membranes, differences in membrane proteins associated with anion transport processes, or differences in the erythrocyte cytoskeleton (Udden, 1994; Udden and Patton, 1994). Humans are most likely to be exposed to BE by the dermal or inhalation routes due to the widespread use of BE in cleaning products. Assessment of the risk resulting from BE use has to account for these routes of exposure and the formation of BAA as the active metabolite. In order to assist in the risk assessment, PBPK models

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were developed that describe the uptake, metabolism and disposition of BE and BAA (Johanson, 1986; Corley et al., 1993, 1994; Shyr et al., 1993). The model of Corley et al. (1993, 1994) is a refinement of Johanson’s model (1986) and consists of two submodels. The first submodel describes the uptake and disposition of BE and consists of the tissue compartments rapidly perfused organs, slowly perfused organs, fat, skin, muscle, gastrointestinal tract, and liver as the metabolizing tissue. The BE submodel allows uptake via the inhalation and dermal routes and in addition provides the possibility of uptake via IV infusion and the gastrointestinal tract in order to validate the model with laboratory data. The second submodel tracks the disposition of BAA in the same tissue compartments, but the kidney was removed from the rapidly perfused organs as separate tissue to allow for the excretion of BAA metabolites. The two submodels are linked together by the metabolism of BE to BAA via a saturable enzymatic pathway catalyzed by alcohol and aldehyde dehydrogenases in the liver. Competing pathways (BE conjugation and BE O-dealkylation) are lumped together and described by an additional enzymatic pathway with Michaelis-Menten kinetics. The model assumes that BAA is bound to proteins in blood and is eliminated by a saturable process in the kidneys. The rate of BAA elimination by the kidneys is described as the sum of glomerular filtration rate of BAA and the acid transport of BAA assuming that no reabsorption occurs. The biochemical constants determined experimentally in the rat were scaled to humans by (body weight)0.7. In the validation process, the model successfully described a wide variety of rat and human data from different laboratories using several routes of administration. BAA was predicted to be formed more rapidly in rats compared with humans, but to be eliminated slower in humans than in rats. In summary, higher maximum concentrations of BAA in blood (Cmax) and also higher areas under the BAA concentration-time curves (AUC) were predicted for rats than for humans, especially as the vapour concentration was increased. For the purpose of dose-response and interspecies extrapolation, BAA-Cmax and BAA-AUC were used as estimates of the internal dose surrogate; Cmax can be related directly to the in vitro haemolysis studies with BAA and is responsive to the dose-rate. The in vitro studies performed (Bartnik et al., 1987; Ghanayem et al., 1987; Ghanayem, 1989; Udden, 1994; Udden and Patton, 1994) suggest that approximately 0.2 mM BAA is required to produce slight haemolysis of rat red blood cells. At about 2 mM BAA nearly complete haemolysis was observed. The model predicts for nose-only exposure that these concentrations are reached in the rat at BE exposure concentrations of about 100 ppm and 800 ppm for 6 h, respectively, which is consistent with observations in vivo (Tyler, 1984; Sabourin et al., 1992).

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For human red blood cells, the minimum BAA-concentration necessary to induce slight haemolysis is about 40 times higher compared with rats, i.e. 8 mM. The model predicts for human nose-only exposure that the Cmax of BAA in blood is slightly lower than the value observed in rats at a BE exposure concentration of about 100 ppm for 6 h and is only about 50 per cent of the BAA rat blood concentration at 800 ppm. In any case, the minimum toxic concentration of approximately 8 mM BAA in human blood is not achieved. The AUC has a time component which is important since haemolysis is not an instantaneous response (Udden, 1994; Udden and Patton, 1994). With respect to the AUCs for BAA, the model predicts that the values for rat and human blood are similar up to a BE exposure concentration of about 500 ppm. Higher BE concentrations cause higher AUCs for BAA in rat blood than in human blood. Thus, the model predicted a BAA-AUC in man at 22 ppm BE vapour exposure that was similar to the BAA-AUC in rats achieved at 25 ppm BE vapour exposure, the established subchronic NOAEL. The simulation of the dermal BE uptake assumed that 10 per cent of the body surface of rats and humans were exposed for 6 h to BE solutions in water (5–100 per cent) and that no losses of BE occurred from the dosing solution. The simulation predicted Cmax blood concentrations for BAA in rats that were highest (about 3 mM) for a 40 per cent BE solution. For humans, BAA-Cmax was predicted to reach about 1.3 mM for the same BEconcentration. Predicted BAA-AUCs were about twofold higher in rats compared with humans. Under these worst-case assumptions, no BE concentration is expected to achieve BAA concentrations in human blood that would cause haemolysis. ECETOC (1994) used the described PBPK model for BE and BAA disposition in combination with mechanistic data obtained by in vitro experiments to recommend an occupational exposure limit for BE: – BAA-AUC was used as the internal dose surrogate and 22 ppm BE vapour was predicted to cause a BAA-AUC in human blood similar to the BAA-AUC in rats exposed to a BE-concentration of 25 ppm (identified as the subchronic rat NOAEL). At 22 ppm BE vapour, the BAA-Cmax (33 µM) is predicted to be several hundredfold below the BAA concentration that causes pre-haemolytic effects in human red blood cells (8 mM). – ECETOC did not use an uncertainty factor for intraspecies extrapolation, since the in vitro studies indicated no increased sensitivity of red blood cells from individuals regarded as susceptible to haemolytic effects such as older persons, persons with hereditary spherocytosis or sickle cell disease (Udden, 1994; Udden and Patton, 1994).

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– An uncertainty factor for time extrapolation (subchronic to chronic exposure) was also not applied, since the red blood cell haemolysis was regarded as a transient phenomenon observed predominantly on the first few days of exposure thus indicating that longer exposure would not have resulted in a lower rat NOAEL. – Although there is some uncertainty about the actual magnitude of the contribution of dermal uptake to the total uptake during BE vapour exposure, ECETOC concluded that even under worst-case conditions the BAA concentrations achieved are not sufficient to cause haemolysis in man and there is no need for the adjustment of the predicted human NOAEL for route. In conclusion, an occupational exposure limit of 20 ppm (8 h TWA) was recommended, also taking into account all other effects that may be associated with BE-exposure. This value is similar to the rat NOAEL of 25 ppm for the most sensitive parameter, i.e. haemolysis, and was derived using scientific data instead of applying default factors to the rat NOAEL, a procedure which would have overpredicted the human risk associated with BE-exposure. Conclusion The use of PBPK models and mechanistic data in risk assessment tends to reduce the uncertainties in comparison with default methodologies by replacing the administered dose with the delivered dose and also tends to reveal uncertainties concealed in default methodologies (Wilson and Cox, 1993). However, there are also limitations in the development of PBPK models. One limitation is that the mechanism of the toxic effect has to be known, otherwise the replacement of the external dose by internal dose surrogates is not possible. In addition, extensive validation of the model is necessary in order to replace default approaches in risk assessment. For the time being, the development of PBPK models appears to be restricted to high production chemicals where the existing data base allows identification of an accepted mechanism of toxic action and validation of the model. Concern has been expressed that the use of point estimates in PBPK modelling instead of ranges of biologically plausible values leads to an increase in the uncertainty (Portier and Kaplan, 1989). However, a recent study from the Delivered Dose Work Group of the American Industrial Health Council came to the conclusion that incorporation of ‘pharmacokinetic information in a risk assessment,…, leads to both a more accurate estimate of risk and a better specification of the true uncertainty’ (Wilson and Cox, 1993). A detailed discussion of the sources of uncertainties is also provided in this reference.

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If the data base is sufficient, PBPK models provide scientific credibility to interspecies extrapolation, extrapolation across routes of administration, extrapolation from high-dose to low-dose and intraspecies extrapolation. Recent concepts link the original tissue dose concept of PBPK models to biologically based tissue response models, thus relating the delivered dose via the mechanism of action to the toxic response and developing integrated biological models (Conolly et al., 1988; Moolgavker et al., 1988; Cohen and Ellwein, 1990; Conolly and Andersen, 1991). Such approaches enable scientists to ask the right questions and to design new mechanistic studies that will lead toward the goal of a scientifically-based risk assessment. Acknowledgement The author thanks Richard A.Corley and the Glycol Ether Panel of the Chemical Manufactures Association for providing data on 2butoxyethanol. References ANDERSEN, M.E. and KRISHNAN, K., 1994, Physiologically based pharmacokinetics and cancer risk assessment, Environ. Health Perspect. Suppl., 102 (Suppl. 1), 103–8. ANDERSEN, M.E., CLEWELL, H.J., GARGAS, M.L., SMITH, F.A. and REITZ, R.H., 1987, Physiologically based pharmacokinetics and the risk assessment process, for methylene chloride, Toxicol. Appl. Pharmacol., 87, 185–205. ARMS, A.D. and TRAVIS, C.C., 1988, Reference physiological parameters in pharmacokinetic modeling, United States Evironmental Protection Agency, EPA/600/6– 88/004, reproduced by US Department of Commerce, National Technical Information Service, Springfield, VA. BARTNIK, F.G., REDDY, A.K., KLECAK, G., ZIMMERMANN, V., HOSTYNEK, J.J. and KUENSTLER, K., 1987, Percutaneous absorption, metabolism, and hemolytic activity of n-butoxyethanol, Fundam. Appl. Toxicol., 8, 59–70. CARFAGNA, M.A. and KEDDERIS, G.L., 1992, Isolated hepatocytes as in vitro models for the biotransformation and toxicity of chemicals in vivo, CIIT Activities, 12, 1–6. COHEN, S.M. and ELLWEIN, L.B., 1990, Proliferative and genotoxic cellular effects in 2-acetylaminofluorene bladder and liver carcinogenesis: biological modeling of the ED01 study, Toxicol. Appl. Pharmacol, 104, 79–93. CONOLLY, R.B. and ANDERSEN, M.E., 1991, Biologically based pharmacodynamic models: tools for toxicological research and risk assessment, Ann. Rev. Pharmacol. Toxicol., 31, 503–23. CONOLLY, R.B., REITZ, R.H., CLEWELL, H.J., III and ANDERSEN, M.E., 1988, Pharmacokinetics, biochemical mechanism and mutational

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accumulation: a comprehensive model for chemical carcinogenesis, Toxicol. Lett., 43, 189–200. CORLEY, R.A., BORMETT, G.A., MARKLEY, B.J. and STACK, C.R., 1993, A physiologically-based pharmacokinetic model for 2-butoxyethanol and its metabolite, 2-butoxyacetic acid, The Toxicologist, 13, 355, 1387A. CORLEY, R.A., BORMETT, G.A. and GHANAYEM, B.I., 1994, Kinetics of 2butoxyethanol and its major metabolite 2-butoxyacetic acid in rats and humans, Toxicol. Appl. Pharmacol., submitted. CSANÁDY, G.A., MENDRALA, A.L., NOLAN, R.J. and FILSER, J.G., 1994, A physiologic pharmacokinetic model for styrene and styrene-7,8-oxide in mouse, rat and man, Arch. Toxicol. (in press). D’SOUZA, R.W., FRANCIS, W.R. and ANDERSEN, M.E., 1988, Physiological model for tissue glutathione depletion and increased resynthesis after ethylene dichloride exposure, J. Pharmacol. Exp. Therap., 245, 563–8. ECETOC, 1985, Technical Report No. 17. The Toxicology of Glycol Ethers and its Relevance to Man: an Updating of ECETOC Technical Report No. 14, Brussels: European Centre for Ecotoxicology and Toxicology of Chemicals. ECETOC, 1994, Special Report No. 7. Butoxyethanol Criteria Document (including a supplement for 2-butoxyethyl acetate), Brussels: European Centre for Eco-toxicology and Toxicology of Chemicals. EEC, 1992, Council Directive 92/32/EEC of 30 April 1992 amending for the 7th time Council Directive 67/548/EEC on the approximation of laws, regulations and administrative provisions relating to the classification, packaging and labelling of substances, Offic. J. Europ. Communities, L 154/1. EEC, 1993a, Council Regulation EEC 793/93 of 23 March 1993 on the evaluation and control of the risks of existing substances, Offic. J. Europ. Communities, L 84. EEC, 1993b, Commission Directive 93/67/EEC of 20 July 1993 laying down the principles for assessment of risks to man and the environment of substances notified in accordance with Council Directive 67/548/EEC, Offic. J. Europ. Communities, L 227/9. FILSER, J.G., 1992, The closed chamber technique—uptake, endogenous production, excretion, steady state kinetics and rates of metabolism of gases and vapours, Arch. Toxicol, 66, 1–10. FREDERICK, C.B., 1993, Limiting the uncertainty in risk assessment by the development of physiologically based pharmacokinetic and pharmacodynamic models, Toxicol. Lett., 68, 159–75. FREDERICK, C.B., POTTER, D.W., CHANG-MATEU, M.I. and ANDERSEN, M. E., 1992, A physiologically based pharmacokinetic and pharmacodynamic model to describe the oral dosing of rats with ethyl acrylate and its implications for risk assessment, Toxicol. Appl. Pharmacol., 114, 246–60. GARGAS, M.L., ANDERSEN, M.E. and CLEWELL, H.J., III, 1986, A physiologically based simulation approach for determining metabolic constants from gas uptake data, Toxicol. Appl. Pharmacol., 86, 341–52. GARGAS, M.L., BURGESS, R.J., VOISARD, D.E., CASON, G.H. and ANDERSEN, M.E., 1989, Partition coefficients of low-weight volatile chemicals in various liquids and tissues, Toxicol. Appl. Pharmacol., 98, 87–99.

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GARGAS, M.L., CLEWELL, H.J. and ANDERSEN, M.E., 1990, Gas uptake inhalation techniques and the rates of metabolism of chloromethanes, chloroethanes and chloroethylenes in the rat, Inhal. Toxicol., 2, 285–309. GHANAYEM, B.I., 1989, Metabolic and cellular basis of 2-butoxyethanol-induced hemolytic anemia in rats and assessment of human risk in vitro, Biochem. Pharmacol., 8, 1679–84. GHANAYEM, B.I., BURKA, L.T. and METTHEWS, H.B., 1987, Metabolic basis of ethylene glycol monobutyl ether (2-butoxyethanol) toxicity: role of alcohol and aldehyde dehydrogenases, J. Pharmacol. Exp. Therap., 242, 222–31. HALLENBECK, W.H., 1993, Quantitative Risk Assessment of Environmental and Occupational Health, Chelsea, MI: Lewis Publisher, Inc. HORTON, V.L., HIGUCHI, M.A. and RICKERT, D.E., 1992, Physiologically based pharmacokinetic model for methanol in rats, monkeys and humans, Toxicol. Appl. Pharmacol., 117, 26–36. JOHANSON, G., 1986, Physiologically-based pharmocokinetic modeling of inhaled 2-butoxyethanol in man, Toxicol. Lett., 34, 23–31. JOHANSON, G. and FILSER, J.G., 1993, A physiologically based pharmacokinetic model for butadiene and its metabolite butadiene monoxide in rat and mouse and its significance for risk extrapolation, Arch. Toxicol., 67, 151–63. KRISHNAN, K., GARGAS, M.L., FENNELL, T.R. and ANDERSEN, M.E., 1992, A physiologically based description of ethylene oxide dosimetry in the rat, Toxicol. Ind. Health, 8, 121–40. LEHMAN, A.J. and FITZHUGH, O.G., 1954, 100-fold margin of safety, Assoc. Food Drug Off. US. Bull., 18, 33–5. LEWIS, S.C., LYNCH, J.R. and NIKIFOROV, A.I., 1990, A new approach to deriving exposure guidelines from ‘no-observed-adverse-effect levels’, Regul. Toxicol. Pharmacol., 11, 314–30. MOOLGAVKER, S.H., DEWANJI, A. and VENZON, D.J., 1988, A stochastic two-stage model for cancer risk assessment. I. The hazard function and the probability of tumor, Risk Anal., 8, 383–92. PARK, C.N. and HAWKINS, N.C., 1993, An overview of cancer risk assessment, Toxicology Methods, 3, 63–86. PORTIER, C.J. and KAPLAN, N.L., 1989, Variability of safe dose estimates when using complicated models of the carcinogenic process. A case study: methylene chloride, Toxicol. Appl. Pharmacol, 13, 533–44. POTTER, D.W. and TRAN, T.-B., 1993, Apparent rates of glutathione turnover in rat tissues, Toxicol. Appl. Pharmacol., 120, 186–92. RAMSEY, J.C. and ANDERSEN, M.E., 1984, A physiologically based description of the inhalation pharmacokinetics of styrene in rats and humans, Toxicol. Appl. Pharmacol., 73, 159–75. REITZ, R.H., MENDRALA, A.L., PARK, C.N., ANDERSEN, M.E. and GUENGERICH, F.P., 1988, Incorporation of in vitro enzyme data into the physiologically based pharmacokinetic model for methylene chloride: implications for risk assessment, Toxicol. Lett., 43, 97–116. REITZ, R.H., MCCROSKY, P.S., PARK, C.N., ANDERSEN, M.E. and GARGAS, M.L., 1990a, Development of a physiologically based pharmacokinetic model for risk assessment with 1,4-dioxane, Toxicol. Appl. Pharmacol, 105, 37–54.

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REITZ, R.H., MANDRELA, A.L., CORLEY, R.A., QUAST, J.F., GARGAS, M. L., ANDERSEN, M.E., STAATS, D.A. and CONOLLY, R.B., 1990b, Estimating the risk of liver cancer associated with human exposures to chloroform using physiologically based pharmacokinetic modeling, Toxicol. Appl. Pharmacol., 105, 443–59. RENWICK, A.G., 1991, Safety factors and establishment of acceptable daily intakes, Food Add. Contam., 8, 135–50. RENWICK, A.G., 1993, Data derived safety for the evaluation of food additives and environmental contaminants, Food Add. Contam., 10, 337–50. SABOURIN, P.J., MEDINSKY, M.A., BIRNBAUM, L.S., GRIFFITH, W.C. and HENDERSON, R.F., 1992, Effect of exposure concentration on the disposition of inhaled butoxyethanol by F-344 rats, Toxicol. Appl. Pharmacol, 114, 232–8. SATO, A. and NAKAJIMA, T, 1979, Partition coefficients of some aromatic hydrocarbons and ketones in water, blood and oil, Br. J. Ind. Med., 36, 231– 4. SHYR, L.J., SABOURIN, P.J., MEDINSKY, M.A., BIRNBAUM, L.S. and HENDERSON, R.F., 1993, Physiologically based modeling of 2-butoxyethanol disposition in rats following different routes of exposure, Environ. Res., 63, 202–18. SILBERGELD, E.G., 1993, Risk assessment: the perspective and experience of US Environmentalists, Environ, Health Persp., 101, 100–4. TARDIF, R., LAPARÉ, S., KRISHNAN, K. and BRODEUR, J., 1993, Physiologically based modeling of the toxicokinetic interaction between toluene and m-xylene in the rat, Toxicol. Appl. Pharmacol, 120, 266–73. TRAVIS, C.C., 1993, Interspecies extrapolation of toxicological data, in Maibach, H. I. (Ed.), CRC Series in Dermatology: Clinical and Basic Science, pp. 387– 410, London: CRC Press. TYLER, T.R., 1984, Acute and subchronic toxicity of ethylene glycol monobutyl ether, Environ. Health Persp., 57, 185–91. UDDEN, M.L., 1994, Hemolysis and decreased deformability of erythrocytes exposed to butoxyacetic acid, a metabolite of 2-butoxyethanol: II. Resistance in red blood cells from humans with potential susceptibility, J. Appl. Toxicol, 14, 97–102. UDDEN, M.L. and PATTON, C.S., 1994, Hemolysis and decreased deformability of erythrocytes exposed to butoxyacetic acid, a metabolite of 2-butoxyethanol: I. Sensitivity in rats and resistance in normal humans, J. Appl. Toxicol, 14, 91– 6. WILSON, A. and Cox, L.A., 1993, Managing Statistical Uncertainties in PBPK Modeling, Denver, CO: Cox Associates.

14 Molecular Approaches to Assess Cancer Risks ALAN S.WRIGHT, J.PAUL ASTON, NICO J.VAN SITTERT and WILLIAM P.WATSON Sittingbourne Research Centre, Sittingbourne

Introduction Carcinogenesis is a complex process which is not yet fully understood. Nevertheless, it is generally accepted that carcinogenesis involves the accumulation of mutations in critical genes: proto-oncogenes and/or tumour suppressor genes. These mutations transform normal cells into ‘initiated’ cells possessing the full complement of genetic changes necessary for malignancy (Figure 14.1). The critical mutations may result from exposures to radiation, to genotoxic chemicals or they may arise ‘spontaneously’ as a consequence of miscoding errors during the normal replication of DNA. Concomitantly, mutations will also accumulate in other genes which, although not critical for cancer per se may, nevertheless, influence cellular character thereby contributing to the multifaceted nature of cancer. The precise nature and number of critical genetic changes required for initiation have not yet been established but will probably vary from case to case. Many researchers envisage a strict temporal sequence of genetic changes in carcinogenesis. However, it is probable that the critical mutations can occur in any sequence and at any time. Indeed, it is clear that one or more of the critical mutations can occur in parental cells. Transmission (inheritance) of these mutations either through the germ line or via somatic cell division increases the susceptibility of the progeny to carcinogens. Furthermore, it is important to note that each of the critical mutations necessary for malignancy may have a different cause. This potential for multiple causation has important implications in risk assessment (vide infra). Fully initiated cells may not automatically proliferate to form tumours. One possible explanation is that the surrounding normal cells restrain the initiated or latent cancer cell by providing essential growth regulators which are no longer produced by the initiated cell. Nevertheless, partially and fully initiated cells have a replicative and/or survival advantage over normal cells. Tissue injury caused by physical trauma, chemical agents or

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Figure 14.1 Schematic representation of the carcinogenic process.

viruses may have a derestraining effect thereby triggering or facilitating the replication of partially or fully initiated cells to form benign tumours or malignant tumours. Increased functional demands may also serve to promote tumour development in affected tissues. It is clear that chemicals which promote tumour development are very important determinants of carcinogenesis. Indeed, promoting agents display a marked tendency for organotropism. Promoter action is, therefore, probably the most important determinant of the site of tumour development. Yet genotoxic chemicals which initiate the carcinogenic process are perhaps viewed with even greater concern. The reasons for this high level of concern hinge mainly on evidence that the mutagenic or initiating actions of genotoxic chemicals are additive, cumulative and essentially irreversible. Furthermore, in contrast to most other classes of toxic chemicals, including promoters operating via cytotoxic mechanisms, there is no theoretical reason or experimental evidence to support the view that mutagenic actions of genotoxic chemicals are thresholded. For these reasons even very low exposures of genotoxic chemicals are viewed with concern. These concerns have focused scientific and regulatory attention on a need to develop sound approaches to manage cancer risks—particularly low level risks associated with low exposures to genotoxic chemicals encountered in the occupational or environmental settings. Indeed, apart from clinical applications, high exposures to genotoxic chemicals cannot be countenanced.

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Management of cancer risks (key requirements) The management of toxicological risks implies a capacity to control exposures within acceptable safety limits. Effective control is, therefore, dependent not only on the qualitative detection and identification of hazardous chemicals but also on a capacity to determine human exposure and to evaluate the health risks. This last requirement necessitates a knowledge of potency, i.e. quantitative human dose-response relationships. In the case of genotoxic chemicals, the relevant data reside in the very low region of the dose-response curve. The concept of acceptable risk is readily accepted when applied to the many classes of toxic chemicals operating by a thresholded mechanism indicative of the virtual absence of risk at sub-threshold doses. Absolute safety margins for genotoxic chemicals cannot be guaranteed (vide supra) leading to the adoption of conservative safety measures. Indeed, cursory analysis might suggest that quantitative risk data are not required for the effective management of cancer risks associated with genotoxic chemicals. Thus, it is generally accepted that human contact with carcinogens should be minimised. Purely qualitative identification of the hazard would permit the design of measures to limit human exposure and minimise carcinogenic impact. Indeed, a purely qualitative indication of genotoxicity can be an absolute deterrent to the development of new products. Nevertheless, certain exposures, e.g. to indigenous genotoxic chemicals and natural food components, are unavoidable. Furthermore, measures to reduce exposures to ‘avoidable’ genotoxic hazards, e.g. certain combustion products, key industrial base chemicals and intermediates, are often difficult and costly. Quantitative risk assessment is needed to prioritise these hazards and, most importantly, to determine safety margins. Certainly a failure to determine carcinogenic potency would lead to uncertainty about the adequacy of safety margins and, probably, to unnecessary measures to further reduce exposures. Thus, despite their additive and cumulative actions, even genotoxic chemicals can pose a negligible health risk. Of course, the definition of a negligible, i.e. acceptable, risk is a socio-political judgement which nevertheless has to be realistic in the case of unavoidable hazards and achievable in the case of avoidable hazards. Detection of genotoxic carcinogens Concerns about genotoxic hazards have provided an incentive for the development of a broad range of rapid tests to detect intrinsic genotoxic activity or potential. The principal aim of these approaches is to predict carcinogenic activity or, more accurately, cancer initiating activity. The most widely used tests are the coupled microsomal-microbial mutation assays developed by Ames et al. (1973). However, such approaches are

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viewed as too remote to be of value in estimating cancer risks. The trend is towards increasingly sensitive and precise technology—particularly generic methods with potential for direct application in humans. Advances in molecular biology have permitted the development of a new generation of point mutation assays based on DNA base mismatch technology (Thilly, 1991; Lu and Hsu, 1992). This technology has a precision far exceeding that of conventional biological methods and a sensitivity permitting direct applications in humans. The full potential of this technology has not yet been realised. However, it seems probable that detection levels will ultimately obviate a need for prior phenotypic selection: paving the way to universal application. Avoidance of phenotypic selection would represent a powerful advantage over existing methodologies by providing a much more direct and reliable route to determining overall background mutation rates and increments due to specific exposures of key relevance to cancer risk assessment (vide infra). The most prospective of the current assays are those designed to detect primary DNA damage. Among these procedures, 32P-post-radiolabelling technology developed by Randerath et al. (1981) to detect DNA adducts is by far the most sensitive. The justification for application of such a prospective approach to detect exposure to genotoxic carcinogens hinges on the causal relationship established between genotoxic activity and cancer. In general, genotoxic character is conferred by possession of a centre(s) of electrophilic reactivity. This reactivity permits the chemical to undergo chemical reactions with nucleophilic centres in the target molecule (DNA). In many instances the electrophilic centre(s) is introduced into an inactive precursor chemical by metabolic activation. Primary products, e.g. DNA adducts, formed when genotoxic chemicals react with DNA are generally promutagenic (or lethal) and their occurrence leads to an increased risk of mutation and cancer. There is no known category of chemical which forms DNA adducts that can be excluded from this generalisation. Not all DNA adducts are strongly promutagenic. However, because electrophiles do not display absolute specificity in their reactions with nucleophiles, the detection of even a weakly promutagenic adduct, e.g. N7-alkyldeoxyguanosine, signals the formation of a more strongly promutagenic adduct, e.g. O6-alkyldeoxyguanosine. If follows that the detection of DNA adducts provides qualitative evidence of (human) exposure to a genotoxic carcinogen.

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Identification of human carcinogens Classical epidemiological approaches Until very recently, epidemiological approaches to detect and identify environmental carcinogens were based exclusively on the analysis of tumour incidence and chromosome aberrations in human populations. However, the endpoints of these biological methods lack the intrinsic resolving power needed to dis criminate between different contributory factors. Indeed, it is only in instances of specific, high and, often, localised exposures that these methods have been effective in identifying specific causative agents. Nevertheless, the results of epidemiological studies indicate that chemicals, which may include both natural and xenobiotic compounds in food, drink or in the local or general environment, play a major and broad role in the aetiology of human cancer. The identification of these chemical factors is a major goal in cancer prevention. In vitro genotoxicity assays In addition to applications in screening prospective chemical products, in vitro genotoxicity assays, particularly the Ames test, provided the first practicable, systematic approach to identify environmental carcinogens. However, this approach places very heavy demands on the time and effort required to fractionate environmental samples and test individual compounds. More importantly, however, like the animal cancer studies these assays complement or have largely supplanted, the approach is not specifically targeted towards identifying and prioritising human hazards. For example, these short-term in vitro test do not provide direct evidence of human exposure or effects. Molecular epidemiology 32P-Post-radiolabelling

technology for the analysis of DNA adducts provides the basis of a very sensitive and generic approach to detect exposures to genotoxic carcinogens. This technology has universal application and can be applied to detect DNA adducts formed in laboratory species or humans during exposures to both known and, as yet, unidentified genotoxic chemicals at the low concentrations encountered in the environment and the workplace. Elucidation of the chemical structures of adducts in human DNA would provide a basis for identifying the causative agents and their sources or origins. This possibility of identifying the chemical initiators of human cancer is an exciting prospect. Unfortunately, however, these adducts are present at very low abundances and this is a major obstacle to identification. Thus, the methods for

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detecting DNA adducts are much more sensitive than the physicochemical methods needed for structural characterisation. A number of strategies have been adopted in attempts to solve this problem. Protein adducts Genotoxic chemicals that react with DNA also react with nucleophilic centres in proteins and may also undergo ‘spontaneous’ and enzymecatalysed reactions with glutathione leading to the excretion of the corresponding mercapturic acids. Insofar as the formation of protein adducts and mercapturic acids reflect the formation of the corresponding DNA adducts, their detection may also furnish evidence of exposure to a genotoxic carcinogen. The potential for reaction of genotoxic chemicals with proteins (and glutathione) is much greater than with DNA. Furthermore, human proteins, e.g. haemoglobin, are available in much larger quantities and are more accessible than human tissue DNA. These advantages have been exploited, particularly in the pioneering work of Ehrenberg’s group, to develop a range of procedures for the qualitative and quantitative analysis of protein adducts (Osterman-Golkar et al., 1976; Calleman et al., 1978; Ehrenberg and Osterman-Golkar, 1980). (For a review of the available methods see Skipper and Naylor, 1991.) The most powerful and generic approach is undoubtedly that developed by Törnqvist et al. (1986a). An initial purification or enrichment step is key to any successful method for the analysis of low levels of organic residues. The amino-groups of the Nterminal valine residues of the α-and β-chains of human haemoglobin are major targets for reaction with a broad range of genotoxic chemicals. Törnqvist achieved selective enrichment of adducted N-terminal valine residues of haemoglobin by devising a modified Edman degradation which resulted in the scission of adducted residues whilst leaving the nonadducted N-terminal valines intact. This procedure provides the basis for identifying the adducting moieties and their quantitation by GC/MS. Applications of this technology have furnished evidence of background exposures to a range of alkylating species. Protein adduct technology has the potential for considerable further refinement. The possibility of using immunoaffinity technology to enrich both known and unidentified protein adducts is currently being explored. DNA adducts and immunoenrichment The need for effective enrichment technology for DNA adducts is even more pressing than for protein adducts. Ideally, the enrichment procedure should be applied at the earliest possible stage of analysis. The procedure

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should be rapid and mild in order to minimise the formation of artefacts. Currently, immunoaffinity technology holds the greatest promise. Immunoenrichment of DNA adducts necessitates antibodies possessing the appropriate specificities and affinities to permit selective binding of adducts at the very low abundances encountered in hydrolysates or enzyme digests of human DNA. The immune system does not normally respond to small molecules per se. However, the system can be induced to produce effective antibodies by immunising with the small molecule (hapten) coupled to a protein. Such treatment induces a spectrum of antibodyproducing cells, each producing a specific antibody. Most of these antibodies recognise various regions (epitopes) of the carrier protein while a few may specifically recognise and bind the small molecule of interest. Suitable antibody-producing cells can be selected and cloned to provide a permanent source of homogenous antibody (monoclonal antibody, Mab). Mabs can be raised against virtually any organic chemical although some lower molecular weight compounds (<150) may pose problems (vide infra). Compound-specific Mabs can be produced which specifically recognise and bind a particular organic chemical. Alternatively, Mabs may recognise a structural feature that is common to a class of chemicals. The prospect of class-specific Mabs permitting, for example, the enrichment of both known and as yet unidentified members of important classes of DNA adducts, e.g. O6-alkyldeoxyguanosine adducts, is particularly exciting. We have conducted a number of collaborative studies with other groups (Dr R.A.Baan, TNO Medical Biological Laboratory, Rijswijk, The Netherlands and Prof. T.Brown, Department of Chemistry, University of Edinburgh, UK) to investigate the potential of immobilised compound- and class-specific Mabs to enrich DNA adducts in an immunoaffinity column mode. Compound-specific Mabs were selected by screening antibodyproducing cells with the small molecule employed in the immunisation process. Class-specific selection may be accomplished by screening with another member of a chemical class, e.g. a lower or higher member of a homologous series. To date Mabs have been raised which recognise a range of O6alkyldeoxyguanosines including the methyl, ethyl, n-propyl and 2hydroxyethyl homologues. Studies are in progress to investigate the utility of these Mabs for the enrichment of O6-alkyldeoxyguanosines, Cooper et al. (1992) have also explored the possibility of developing immunoaffinity systems for the enrichment of a range of O6-alkylated purines and have reported a 107–108-fold purification of O6-methyldeoxyguanosine. Ring-opened N7-deoxyguanosine (N7-dG) adducts possess unusual structural features which distinguish them from all other classes of adducts and from their parent nucleoside (dG). These ring-opened adducts may therefore prove to be particularly effective haptens in approaches to develop generic immunoenrichment methods for DNA adducts. N7-dG

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adducts are not viewed as particularly promutagenic. Nevertheless the N7atom of dG residues in DNA is a major target for adduction and the detection of N7-dG adducts signals the production of adducts at other (more critical) sites in DNA (vide supra). In certain instances, a class of adducting moieties may possess a common structural feature that can be exploited for immunoenrichment. For example, a Mab has been raised against the major DNA adduct of benzo(a) pyrene (r-7,t-8,t-9-trihydroxy-c-10-(N2-deoxyguanosylphosphate)-7,8,9, 10-tetrahydrobenzo(a)pyrene) in a collaborative study with Dr Baan’s group. This Mab recognises DNA adducts formed by a broad range of polycyclic aromatic hydrocarbons (PCAs) including benzo(a)pyrene (BP), chrysene, benz(a)anthracene, 5-methylchrysene, picene and dibenz(a,h) anthracene. It seems probable that this Mab recognises the common trihydric alcohol structure produced when the reactive diol epoxides of each of these polycyclic compounds reacts with nucleophilic centres in DNA or other macromolecules. The fact that the Mab does not bind the corresponding fluoranthene adduct is consistent with the spatial environment of the hydroxyl groups in fluoranthene-DNA adducts which is completely different from those generated from the other PCAs employed in this study. The performance of the Mab raised against the major BP-DNA adduct in the enrichment of PCA-DNA adducts is being evaluated using the immobilised Mab coupled to cyanogen bromide-activated Sepharose 4B. Results obtained to date demonstrate that the immobilised Mab selectively adsorbs the major BP-DNA adduct from DNA hydrolysates at abundances below 1 adduct per 109 nucleotide units. Results with the other PCA-DNA adducts are not yet available. However, the results obtained with the major BP-DNA adduct underlines the potential of immunoenrichment technology in the qualitative and quantitative analysis of adducts. Furthermore, such results provide an incentive to pursue the development of class-specific antibodies in order to permit or facilitate the identification of the chemical initiators of human cancer. Mercapturic acids Qualitative analysis of mercapturic acids also provides a basis for identifying human exposures to genotoxic chemicals (vide supra). However, the available analytical procedures are complex and tend to lack specificity and sensitivity. During the last dozen years we have undertaken a number of studies aimed at developing compound- and class-specific antibodies to facilitate the analysis of mercapturic acids. Conventional approaches to generate antibodies to low molecular weight (MW) organic chemicals involves the covalent attachment of the small molecule (hapten) to a strongly antigenic protein, e.g. keyhole limpet

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haemocyanin or bovine serum albumin, for the immunisation of mice. This strategy is usually effective in the case of strongly antigenic haptens, e.g. aromatic nitro compounds of PCA-DNA adducts. However, all of our attempts to use this approach to generate antibodies against relatively low MW and weakly antigenic mercapturic acids, e.g. S-(2-hydroxyethyl)-Nacetylcysteine, failed. Antibodies were generated but these were directed against the strongly antigenic carrier protein(s). Covalent binding to macromolecules is believed to provide the basis of allergic responses, e.g. skin sensitisation reactions, to small molecules and, possibly, a basis for the induction of auto-immune responses. Thus, the binding of the small molecule transforms normal proteins into ‘foreign’ proteins which trigger an immune response. Recently we have employed this principle in an attempt to direct the immune response specifically against mercapturic acid haptens by immunising mice with the haptens bound to a non-antigenic carrier protein, i.e. mouse serum albumin. Preliminary results indicate that this tactic has been successful. Overall the treatment induced fewer antibody-producing cells. However, the antibodies that were generated show high affinities and specificity toward model mercapturic acids including S-(2-hydroxyethyl) and S-phenylmercapturic acid. Studies are in progress to investigate the performance of these antibodies in an immunoenrichment mode. The preliminary results of our studies using non-antigenic protein carriers are very encouraging and have provided fresh insights which may assist in directing immune responses against the specific structural features of interest. Improvements in our ability to tailor the antibody will prove extremely valuable in optimising the properties of antibodies to meet specific needs, e.g. to enrich DNA, protein or mercapturic acid adducts for application in identifying the chemical initiators of human cancer and quantifying exposures to these agents. Cancer risk assessment Human exposure monitoring (determination of dose) The assessment of cancer risks posed by exposure to genotoxic chemicals has two components: determination of the dose and determination of the effect (increment in cancer incidence) caused by that dose. The introduction of the target dose concept by Ehrenberg in the early 1970s has provided the key to modern strategies to assess genotoxic risks (Ehrenberg, 1974, 1979; Ehrenberg et al., 1974). This new dose concept was developed to provide a measure of the critical dose, i.e. the dose of the ultimate genotoxic agent(s) penetrating to DNA. Target dose is much more relevant to risk assessment than is exposure dose. The determination of target dose automatically

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compensates for individual or species differences in the operation of metabolic and biokinetic factors that control the quantitative (and qualitative) relationships between the exposure and the dose of the ultimate toxicant delivered to the target. Measurements of target dose may be applied to improve the extrapolation of risk data from experimental models to humans and may also provide improved definitions of risks to individuals. The determination of target dose in humans may be viewed, therefore, as an approach towards direct risk monitoring as well as a more relevant approach to monitor human exposures to genotoxic chemicals. Determination of target dose The determination of target dose raises numerous technical and theoretical problems. Target dose can be determined by measuring primary products, e.g. DNA adducts, formed when genotoxic agents react with DNA. The kinetics of formation and decay of these adducts must also be determined (vide infra) in order to transform measurements of amounts of adducts into estimates of target dose. Human tissue DNA is not readily accessible for monitoring purposes: surrogate dose monitors are required. There are numerous possibilities including the determination of adducts in white blood cell DNA or of the corresponding adducts in the haemoglobin of circulating erythrocytes. Such indirect approaches require validation. Haemoglobin is the most extensively studied surrogate, not only because of its accessibility and relative abundance but also because of the relative stability of haemoglobin adducts and the longevity of erythrocytes which permit retrospective estimates of dose received by the erythrocytes over a period of about 4 months. Current evidence indicates that all electrophiles that undergo covalent reactions with DNA also react with haemoglobin. Furthermore the amounts of haemoglobin adducts are quantitatively related to the rates of formation of DNA adducts in the tissues. However the proportional relationships between the doses delivered to tissue DNA and to haemoglobin or to any other surrogate will vary from chemical to chemical and will have to be established using experimental models. Measurement of haemoglobin adducts Genotoxic chemicals undergo covalent reactions with a variety of nucleophilic centres in haemoglobin including the sulphydryl group of cysteine, the N1 and N3 atoms of histidine and the amino groups of Nterminal valine residues. Ehrenberg’s group (Osterman-Golkar et al., 1976; Calleman et al, 1978; Törnqvist et al., 1986a) has pioneered the development of methods to detect, identify and quantify adducts formed at each of these centres. A review of these and methods for the analysis of

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‘labile’ adducts formed, for example, during exposure to aromatic amines (Green et al., 1984; Albrecht and Neumann, 1985) is beyond the scope of this paper (for reviews see Farmer, 1991; Skipper and Naylor, 1991). However, probably the most powerful and valuable approach was developed by Törnqvist et al. (1986a, b) who showed that adducts with the N-terminal valine residues of haemoglobin could be specifically enriched by scission in a modified Edman reaction followed by extraction. This enrichment procedure greatly facilitates sample analysis by GC/MS. Immunoassays are also being introduced as alternatives to physicochemical methods for the determination of protein adducts (Wraith et al., 1988). However, as in the case of DNA adducts, the biggest impact of immunotechnology on the analysis of protein adducts will probably be in the immunoenrichment of low levels of adducts for analysis by physicochemical methods. Determination of biological effects Tumour incidence The determination of target dose is essential for assessing cancer risks posed by low-level exposures to genotoxic chemicals. The other requisite is know ledge of the human dose-carcinogenic response relationships in the low-dose range. The lack of intrinsic resolving power of classical epidemiological methods (vide supra) prevents effective applications to detect small carcinogenic effects associated with low exposures to any particular genotoxic chemical. Furthermore, the detection limits of animal cancer studies fall short of ‘acceptable’ risk limits by three to four orders of magnitude (Wright, 1991). This poor sensitivity compels the use of high test doses in order to ensure that significant carcinogens do not go undetected. However, it is generally accepted that high doses of chemicals may induce tumours by non-specific mechanisms, e.g. via tissue injury and compensatory cell proliferation, that do not operate at low doses (Ames, 1989; Wright, 1991). Many, if not all genotoxic chemicals induce cell injury at high (thresholded) doses. Clearly, extrapolation of such high dose risk data to the relevant low-dose range may, at the very least, lead to a gross overestimation of risk. Determination of mutagenic potency In considering the impact that a low-level exposure to a genotoxic chemical may have on cancer incidence, it is reasonable to suggest that the mutagenic propensity of the chemical, although of a low order, would nevertheless be the overriding risk factor. Thus, it is probable that any

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intrinsic promoter activity of the chemical arising, for example as a consequence of cell injury, would be negligible at low exposures— particularly when viewed in the context of the overall promoter pressure exerted on the populations at risk. Thus, it is unlikely that the added cancer risk would be greater than and may approximate to the increment in critical mutations caused by the specific exposure (Figure 14.1). Experimental studies indicate that all known categories of genotoxic agents ranging from methylating agents to polycyclic aromatic compounds can induce the critical mutations leading to malignancy (Figure 14.1). Furthermore, at low doses, the possibility that exposure to a particular genotoxic chemical would induce more than one of the critical mutations in any particular cell is extremely remote. Indeed it is probable that each critical mutation is induced by a different agent or mechanism, i.e. chemical, radiation or ‘spontaneous’. In this sense, each critical mutation would have equal status, i.e. no single event would be any more or any less critical than any other to the final outcome. Accordingly, the increment in cancer risk would equate with the increment in any decisive, e.g. oncogeneactivating, mutation in any critical gene. The induction of such mutations is almost certainly a direct function of overall mutagenic activity of the chemical, i.e. linked to the number of mutational events rather than the type of mutations induced by a given dose of the chemical. At low exposures, therefore, the increment in cancer incidence due to a specific genotoxic agent would approximate to the small increase in the total mutational load caused by the exposure, i.e. relative to the overall background level of mutations due to all causes, multiplied by the overall cancer incidence in the population at risk. (The latter function introduces a measure of the net impact of promoter and anti-promoter pressure acting on initiated cells in the population at risk.) Of course risks may also be calculated on the basis of specific tumours and specific tissues. The determination of small increments in mutation associated with lowlevel exposures to genotoxic chemicals in human populations presents enormous technical problems not least due to the much larger and variable background of mutations due to all causes. Increasing the sensitivity of mutation assays per se (vide supra) is unlikely to improve the situation. High resolving power is also needed to discriminate between effects due to different contributory factors. Nevertheless, while direct approaches to assess absolute cancer risks posed by such low-level exposures may elude us we can nevertheless begin to determine relative cancer risks and prioritise genotoxic chemicals on the basis of experimental determinations of mutagenic potency and estimates of target doses resulting from environmental or occupational exposures. Thus, according to the foregoing the relative cancer risk posed by a low exposure to a genotoxic agent would approximate to the number of mutations induced per unit target dose×estimated human target dose. Once relative cancer risks have been

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established, the determination of the ‘absolute’ risks for one genotoxic chemical would permit calculation of the ‘absolute’ risks for the others. Such ‘absolute’ risks will nevertheless vary from population to population dependent upon variations in promoter pressure. Determination of ‘absolute’ cancer risks Increments in human mutation caused by low-level exposures to genotoxic chemicals are essential for risk estimation but cannot be determined directly (vide supra). Such increments may be estimated using experimental models. However, unless the variations in background are of a very low order, it is unlikely that even the most sensitive of the emerging mutation assays will permit the measurement of small increments of mutation at low, e.g. environmental, exposures. Extrapolation to low doses will be required and must necessarily be conservative, i.e. linear extrapolation to the origin. In addition to ‘high’ dose-low dose extrapolation, it will be necessary to apply corrections for differences between the model and humans in the operation of systemic factors that govern the relationships between exposure and mutagenic effect. Estimates of target dose in the human population at risk and in the experimental model compensate for differences in metabolic and biokinetic factors that determine the relationships between exposure and the critical dose. In effect, the determination of target dose provides a measure of the rates of formation of the key (primary and critical) chemical lesions leading to mutation. The final stage in translating the experimentally-determined risk data to humans is to apply corrections for systemic factors that determine the progression of the key lesions into mutations. The equivalent radiation dose concept The principal systemic factors determining the progression of key chemical lesions in DNA into mutations are the rates and fidelities of DNA repair and replication (Wright et al., 1988). Ehrenberg and co-workers have suggested that the repair of primary DNA damage induced by low doses of radiation may be proportionate to that induced by low doses of genotoxic chemicals. They have further suggested that the determination of the relative mutagenic effectiveness or potencies of radiation and any particular genotoxic chemical may be of value in correcting for species differences in factors determining the progression of primary DNA damage into mutations. The model proposed by Ehrenberg (1980) is based on the determination of the dose-response curves for the induction of the same mutation in the same experimental system by low target doses of the test chemical and acute -radiation. A consistent ratio between the two curves,

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which need not be linear, would indicate proportionality over the low dose range of interest and would permit the mutagenic potency of the chemical to be expressed in terms of radiation equivalents, i.e. the number of rads giving the same response or risk as a unit of chemical dose (expressed in terms of target dose, e.g. millimolar hour, mM h). The significance of such radiation dose-equivalents hinges on their possible extrapolative value. Thus, in order to be useful in assisting the translation of experimentallydetermined mutagenicity data, the rad-equivalence value for the test chemical must have a similar numerical value in both the test system used to determine mutagenicity and in humans. However, it is improbable that rad-equivalence values can be directly determined in humans. Rad-equivalence values for the induction of mutations have been determined for a number of intrinsically reactive monofunctional alkylating agents using a wide range of genetic endpoints in a variety of biological systems including bacteria, plants and mammalian species—the latter, mainly in vitro (Ehrenberg et al., 1974; Ehrenberg, 1976, 1979; Calleman, 1984; Kolman et al., 1989). The rad-equivalence value for a given alkylating agent was approximately the same (within a factor of two) in each of the test systems. On the basis of such evidence Calleman et al. (1978) concluded that there was no reason to presume that a value for radequivalence established in these disparate systems would differ in humans. The best studied example is ethylene oxide. Currently a conjoint programme is underway at the Universities of Stockholm and Leiden to determine radequivalence values in rodents in vivo using a variety of endpoints including the clonal HGPRT mutation assay and induction of pre-neoplastic nodules in rat liver. Preliminary findings are encouraging (Ehrenberg, personal communication). Demonstration of their extrapolative value would justify application of rad-equivalence values to compute small increments in mutation induced in humans by low exposures (determined as target dose) to genotoxic chemicals. The determination of these increments is the basis of the risk model (vide supra) in which the increment in cancer risk due to a particular chemical in a population is viewed as approximating to the increment in mutation induced by the chemical. However, the fact that risk coefficients have not yet been established for radiation-induced mutations in humans precludes applications of rad-equivalence values to estimate mutational risks, e.g. small increments in mutation, in humans. Of course, Ehrenberg realised that a genotoxic chemical(s) may prove to be superior to radiation as a reference standard for estimating mutational risks. Indeed, the use of chemicals that are representative of classes of genotoxic chemicals, repair pathways, etc. is envisaged in this developing ‘equivalence’ strategy (Törnqvist and Osterman-Golkar, 1991). However, at the time the original strategy was formulated there was and still is no reliable data relating low level exposure to a genotoxic chemical and the attendant risk of mutation

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or cancer in human populations. In contrast, risk coefficients have been established for the induction of cancer by low levels of -radiation in human populations within, approximately, a factor of 2. According to the basic presumptions (vide supra), the increased cancer risks in a human population caused by low-level exposures to radiation or to a genotoxic chemical are predominantly due to the mutagenic propensities of these genotoxic factors. It would, therefore, seem reasonable to suggest that cells that had been initiated by exposure to radiation or to genotoxic chemicals or combinations of factors (vide supra) would, nevertheless, be subject to the same general promoter and modulating influences acting on the population. On this basis, experimentally-determined rad-equivalence values for genotoxic chemicals may be used to convert target doses of the chemicals (determined in human populations receiving low-level exposures) into the equivalent doses of radiation (for the induction of mutation). The respective cancer risks associated with any particular monitored target dose of a genotoxic chemical may then be obtained by direct reference to the human cancer risk coefficients established for radiation. References ALBRECHT, W. and NEUMANN, H.-G., 1985, Biomonitoring of aniline and nitrobenzene. Haemoglobin binding in rats and analysis of adducts, Arch. Toxicol, 57, 1–5. AMES, B.N., 1989, Environmental pollution and the causes of human cancer: six errors, in DeVita, V.T., Jr, Hellman, S. and Rosenberg, S.A. (Eds) Important Advances in Oncology, pp. 237–47, Philadelphia, PA: Lippincott. AMES, B.N., DURSTON, W.E., YAMASAKI, E. and LEE, F.E., 1973, Carcinogens are mutagens: a simple test system combining liver homogenates for activation and bacteria for detection, Proc. Nat. Acad. Sci. USA, 70, 2281–5. CALLEMAN, C.J., 1984, Haemoglobin as a dose monitor and its application to the risk estimation of ethylene oxide, PhD Thesis, p. 25, Stockholm: University of Stockholm. CALLEMAN, C.J., EHRENBERG, L., JANSSON, B., OSTERMAN-GOLKAR, S., SEGERBÄCK, D., SVENSSON, K. and WACHTMEISTER, C.A., 1978, Monitor ing and risk assessment by means of alkyl groups in haemoglobin in persons occupationally exposed to ethylene oxide, J. Environm. Pathol. Tox., 2, 427–42. COOPER, D.P., GRIFFIN, K.A. and POVEY, A.C., 1992, Immunoaffinity purification combined with 32P-postlabelling for the detection of O6methylguanine in DNA from human tissues, Carcinogenesis, 13, 469–75. EHRENBERG, L., 1974, Genotoxicity of environmental chemicals, Acta. Biol. Yugosl., Ser. F Genetika, 6, 367–98. EHRENBERG, L., 1976, Methods of Comparing Effects of Radiation and Chemicals, Brighton IAEA Consultant Meeting.

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EHRENBERG, L., 1979, Risk assessment of ethylene oxide and other compounds, in McElheny, V.K. and Abrahamson, S. (Eds) Assessing Chemical Mutagens: The Risk to Humans (Banbury Report 1), pp. 157–90, Cold Spring Harbour, New York, CSH Press. EHRENBERG, L., 1980, Purposes and methods of comparing risks of radiation and chemicals, in Radiobiological Equivalents of Chemical Pollutants, pp. 11– 36, Vienna: International Atomic Energy Agency. EHRENBERG, L., HIESCHE, K.D., OSTERMAN-GOLKAR, S. and WENNBERG, I., 1974, Evaluation of genetic risks of alkylating agents: tissue doses in the mouse from air contaminated with ethylene oxide, Mutat. Res., 24, 83–103. EHRENBERG, L. and OSTERMAN-GOLKAR, S., 1980, Alkylation of macromolecules for detecting mutagenic agents, Teratogen., Carcinogen. Mutagen., 1, 105–27. FARMER, P.B., 1991, Analytical approaches for the determination of proteincarcinogen adducts using mass spectrometry, in Groopman, J.D. and Skipper, P. L. (Eds) Molecular Dosimetry and Human Cancer: Analytical, Epidemiological and Social Considerations, pp. 189–210, Boca Raton: CRC Press. GREEN, L.C., SKIPPER, P.L., TURESKY, R.J., BRYANT, M.S. and TANNENBAUM, S.R., 1984, In vivo dosimetry of 4-aminobiphenyl in rats via a cysteine adduct in haemoglobin, Cancer Res., 44, 4254–9. KOLMAN, A., NASLUND, M., OSTERMAN-GOLKAR, S., SCALIA-TOMBA, G.P. and MEYER, A.L., 1989, Comparative studies of in vitro transformation by ethylene oxide and gamma-radiation of cells, Mutagenesis, 4, 58–61. Lu, A-L. and Hsu, I-C., 1992, Detection of single DNA base mutations with mismatch repair enzymes, Genomics, 14, 249–55. OSTERMAN-GOLKAR, S., EHRENBERG, L., SEGERBÄCK, D. and HALLSTROM, I., 1976, Evaluation of genetic risks of alkylating agents. II. Haemoglobin as a dose monitor, Mutat. Res., 34, 1–10. RANDERATH, K., REDDY, M.V. and GUPTA, R.C., 1981, 32P-labeling test for DNA damage, Proc. Natl Acad. Sci. USA, 78, 6126–9. SKIPPER, P. L and NAYLOR, S., 1991, Mass spectrometric analysis of proteincarcinogen adducts, in Garner, R.C, Farmer, P.B., Steel, G.T. and Wright, A.S. (Eds) Human Carcinogen Exposure: Biomonitoring and Risk Assessment, pp. 61–8, Oxford: Oxford University Press. THILLY, W.G., 1991, Mutational spectrometry: opportunities and limitations in human risk assessment, in Garner, R.C, Farmer, P.B., Steel, G.T. and Wright, A. S. (Eds) Human Carcinogen Exposure: Biomonitoring and Risk Assessment, pp. 127–33, Oxford: Oxford University Press. TÖRNQVIST, M. and OSTERMAN-GOLKAR, S., 1991, Monitoring of in vivo dose by macromolecular adducts: usefulness in risk estimation, in Groopman, J.D. and Skipper, P.L. (Eds) Molecular Dosimetry and Human Cancer: Analytical, Epidemi ological and Social Considerations, pp. 89–102, Boca Raton: CRC Press. TÖRNQVIST, M., MOWRER, J., JENSEN, S. and EHRENBERG, L., 1986a, Monitoring of environmental cancer initiators through haemoglobin adducts by a modified Edman degradation method, Analyt. Biochem., 154, 255–66.

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TÖRNQVIST, M., OSTERMAN-GOLKAR, S., KAUTIAINEN, A., JENSEN, S., FARMER P.B. and EHRENBERG, L., 1986b, Tissue doses of ethylene oxide in cigarette smokers determined from adduct levels in haemoglobin, Carcinogenesis, 7, 1519–21. WRAITH, M.J., WATSON, W.P., EADSFORTH, C.V., VAN SITTERT, N.J., TÖRNQVIST, M. and WRIGHT, A.S., 1988, An immunoassay for monitoring human exposure to ethylene oxide in Bartsch, H., Hemminki, K., and O’Neill, I.K. (Eds) Methods for Detecting DNA Damaging Agents in Humans: Applications in Cancer Epidemiology and Prevention, IARC Scientific Publications No. 89, pp. 271–4, Lyon, France: International Agency for Research on Cancer. WRIGHT, A.S., 1991, Emerging strategies for the determination of human carcinogens, detection, identification, exposure monitoring and risk evaluation, in Garner, R.C, Farmer, P.B., Steel, G.T. and Wright, A.S. (Eds) Human Carcinogen Exposure: Biomonitoring and Risk Assessment, pp. 3–23, Oxford: Oxford University Press. WRIGHT, A.S., BRADSHAW, T.K. and WATSON, W.P., 1988, Prospective detection and assessment of genotoxic hazards: a critical appreciation of the contribution of L.G.Ehrenberg, in Bartsch, H., Hemminki, K. and O’Neill, I.K. (Eds) Methods for Detecting DNA Damaging Agents in Humans: Applications in Cancer Epidemiology and Prevention, IARC Scientific Publications No. 89, pp. 237–47, Lyon, France: International Agency for Research on Cancer.

15 Evaluation of Toxicity to the Immune System HANS-WERNER VOHR Bayer AG, Wuppertal

Introduction A number of years ago the new field of immunotoxicology was established. Very early on, calls were heard from various sides demanding that the development of new chemicals should also take into account the influence of these substances on the immune system (Dean, 1979; Luster et al., 1988; Trizio, et al., 1988). These demands led on the one hand to the initiation of a number of investigations and collaborative studies (ICICIS; BGA; USNTP), on the other hand to thoughts by the authorities and industry on the introduction of guidelines (US-EPA, 1982, 1990; Sjoblad, 1988; ECETOC, 1990; UK-DOH, 1991; Hinton, 1992; OECD, 1992ab. If we define immunotoxicology as the science of adverse effects of substances on the immune system we can say further that these side-effects can lead to either immunopotentiation or immunosuppression. The former can lead to induction of autoimmune reactions and to Type I-IV hypersensitivity reactions, the latter to reduced resistance to infection, development of cancer and also to autoimmune phenomena. On the basis of this definition, immunotoxicological investigations have already been carried out for years during the development of substances; namely with respect to DTH reactions (Type IV) in the guinea pig (Bühler, 1965; Magnusson and Kligman, 1969). As an alternative to these tests in the guinea pig, the so-called local lymph node assay (LLNA) in the mouse according to Kimber et al. (1989) was developed and validated and has meanwhile been adopted as alternative test in the OECD guidelines (OECD, 1992a, b; Botham et al., 1991). The development or selection of suitable tests for immunotoxicological screening and thus for incorporation in guidelines presents considerable problems. Most of the tests which have been proposed for immunotoxicological investigations and most knowledge and experience in immunology are based on mouse models. The standard animal in the early phase of toxicological testing, however, is the rat. Transference of the tests is not always easy, partly because of lack of suitable reagents.

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The next problem is to find which tests can suitably be used—simply and without having to treat additional animals—for reliable identification of interactions with the immune system. Another question which has not yet been solved is relating to the dosages and the changes in immunological parameters which are still tolerable and at which times these should be determined. National and international collaborative studies Most of the guideline drafts favour a two-tiered to three-tiered approach for the screening of immunotoxic side effects, with the greatest disparities in respect of the proposals for tier I tests, ranging from just organ weights and a little more emphasis on the histology of the lymphatic organs, to a series of elaborate, supplementary function tests including, in some cases, satellite groups. Any discussion about basic tests is hampered by a lack of data from routine toxicological and/or epidemiological studies although a few years ago a number of collaborative studies were initiated, namely: ICICIS (international), US-NTP (international, USA (two)), BGA (Federal German Health Office, Germany), GEVI (France). The aim of all these studies is to check on different histological parameters and functional tests for their possibility to flag an immunotoxic potential of a compound in a routine 14 or 28 day study (rats) in an interlaboratory trial. Two immunosuppressive standards (azathioprine and/or cyclosporin A) have been used so far. Although the experimental phase is finished the evaluation of the data is still underway. Table 15.1 summarises the immunotoxicological experiments and collaborative studies currently in progress. With a few exceptions all investigations are based on a 28-day gavage study in rats. These basic tests were supplemented by extended histopathology and functional tests. ICICIS The first substance to be investigated in the international collaborative study was azathioprine. However this first attempt suffered from lack of harmonisation between the models used by the 28 participants world-wide. This made the comparability and the evaluation of the results particularly problematic. The second substance which was tested by ICICIS in a more restrictive design with slightly fewer participants was cyclosporin A. The experimental section as well as the first evaluation of the results has now been completed. As the final evaluation—including statistics—of ICICIS is likely to take some time (years?) it will not be discussed further at this point.

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Table 15.1 Immunotoxicology collaborative studies

Fischer 344 study (Kimber-White) An other interesting approach was done by Kimber-White and colleagues. In this study the duration of treatment was 14 days, Fischer 344 rats were used as experimental animals. Apart from the usual parameters the plaque assay (PFCA), mitogenic stimulation (ConA, LPS) and NK activity were measured.

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No effects were found with respect to organ weights or cell counts analysis but there were marked effects in the PFCA. Here there was a good correlation between the laboratories and clearly dose-dependent effects were seen. Mitogenic stimulation was not and NK activity only slightly affected. Results of the BGA collaborative study In order to put the discussion on a somewhat sounder footing it is imperative to test the various models for detection of immunotoxicological potential in practice. For this reason Bayer AG is taking part in a collaborative trial initiated by the German Federal Health Office in Berlin (BGA study). In this collaborative study standard immunotoxic substances are investigated in parallel under restrictive conditions in several laboratories. On the other hand Bayer AG has introduced a set of functional immunological tests into the routine toxicological testing of agrochemicals in rats in order to test the informative value of these parameters in practice (Vohr, 1995). In the course of this collaborative study cyclosporin A was investigated as the first substance in a very well harmonised design. One reason for choosing cyclosporin A was to permit comparison with the ICICIS results. The study was based on OECD guideline 407 which was supplemented by a number of histopathological, haematological, clinical-chemical and functional parameters. The final evaluation, including pathology and statistics will take a few more months. Nevertheless a first look at the data showed that there were no marked effects with respect to either organ weights (lymphatic organs) or blood cells. On the other hand dose-related effects were found for many parameters in the functional tests. Examples are shown of these parameters and their changes at the various dosages. Both sexes show dose-related changes from the lowest dosage (1 mg kg−1) upwards with respect to the surface markers of the immunocompetent cells. The PFCA and the measurements of IgA antibodies in serum also proved to be sensitive parameters. In immunotoxic investigations based on data obtained in rats treated with test substances (28-day test) we found that secondary influences often occur and that occasionally one single parameter is changed at the highest dosage. Apart from these non-specific effects, dose-related reactions were observed from the lowest or middle dose upwards (for example in the case of cytostatic substances). Such genuinely immunotoxic compounds were reliably identified by surface markers (like CD4/CD45R or PanB) and changes in the serum Ig-titres. Changes in other parameters such as cell counts and macrophage activity verified these findings.

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Findings of the US-NTP study (Luster) The US-NTP study investigated 51 substances, 35 of which were declared immunotoxic, in a comprehensive test battery in mice for changes in functional parameters after 28-day administration of the substances. The correlations of each of these parameters with the given classification and with the results of host-resistance studies were calculated. The correlations after combinations of individual tests were also calculated (Luster et al., 1992, 1993). The conclusion drawn from these investigations was that the immunotoxic potential of a substance can be determined relatively reliably by combination of 2–3 specific tests. The most powerful tests proposed by Luster et al. for such a combination include surface markers, NK test and PFCA. Serum titres of Ig —particularly IgA—were unfortunately not determined. With regard to the correlation with host resistance (HR) results it was found that if effects were shown in the HR model there were always effects on functional parameters, too. There were, however, also cases in which there were effects in the functional tests although the HR studies were negative. Although these investigations were carried out on mice and the choice and classification of the substances are not entirely undisputed, these findings are nevertheless confirmed by our own experience. Discussion and prospect It is undoubtedly too early to make any judgement. However, it appears that apart from the histology—particularly the immuno-histology—a few additional parameters such as analysis of surface markers of subpopulations and serum titre assays of IgG, IgM and IgA are sufficient as screening indicators to show the possible immunotoxic potential of a substance. One of these is the PFCA, which presupposes, however, that satellite groups are used or that the authorities accept injection of SRBC as GLP treatment. Positive findings in a combination of these tests should then occasion further elucidation of the immunotoxic potential. In summary it can be concluded from the currently available results of the collaborative studies that the following criteria must be fulfilled if a substance is to be characterised as possibly immunotoxic. The substance must: 1. Induce significant dose-related changes in one of the effective tests listed above, or 2. induce significant changes in the highest dose group in a combination of 2–3 of these tests.

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But also findings in lymphoid organs which are remarkable with respect to quality, severity or quantity of changes should be of sufficient relevance to warrant further assessment. Finally I would like to point out the problems involved in the evaluation of immunological changes. • In contrast to other data obtained in routine toxicology, historical data and routine experience with respect to the functional tests are so far almost entirely lacking. On account of this paucity of data a decision as to the right effect level for the functional tests can only be made on the basis of subjective judgement at present. The discussion of uniform criteria is only just beginning. • If effects only occur secondarily, e.g. on account of inflammatory processes, they are not specifically immunotoxic. It must be discussed whether such effects on the immune system, which is only doing its normal job in these cases, can be classed as immunotoxic at all. • The immune system can show an ‘oscillating’ response to a substance. A substance may have a stimulating effect in a low dose range and an inhibitory effect in a high range or vice versa. Such reactions are also dose-related effects. • Most immunotoxic investigations have so far used known immunosuppressive drugs. There are, however, little data—particularly from rat studies— on immunostimulant substances. Since the other unwanted side-effect apart from immunosuppression is immune potentiation, future collaborative investigations must urgently concentrate on such substances. Before this, no final evaluation and thus no recommendations on relevant tests can be made. For pharmaceuticals (Hinton, 1992), pesticides (US-EPA, 1990, 1993) and veterinary medicinal chemicals (EEC, 1991) final drafts or notes for guidance for the screening of the immunotoxic potential of a compound already exist. These draft proposals for immunotoxicity parameters for incorporation into new guidelines are shown in Tables 15.2 (FDA) and 15.3 (US-EPA). For industrial chemicals advanced screening of lymphoid organs also with respect to functional parameters had been expected to be incorporated into the adopted OECD guideline No. 407 (1992). But recommendations made by van Loveren and Vos (1992) have not yet been taken into consideration for this update of OECD guideline 407. The proposal of these authors recommended more histopathology (gut associated lymphoid tissue), measurement of serum immunoglobulins, bone marrow cellularity, cytofluorimetry of spleen cells and measurement of NK cell activity. A task force ‘Immunotoxicology’ initiated by ECETOC has put forward proposals on hazard identification and risk assessment of immunotoxic

Abbreviations: CBC=complete blood cell count; WBC=white blood cell count; Ig=Immunoglobin; NK=natural killer; IL-2=interleukin 2; SRBC=sheep red blood cells; TNP-LPS=trinitrophenol lipopolysaccharide. * Recommended for inclusion in basic testing.

Table 15.2 Summary of immunotoxicity testing recommendations for direct food additives

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potential of a compound on the basis of the routine 28 day treatment of rats (Basketter et al., 1994; ECETOC, 1994). A central point in these

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Table 15.3 Proposed amendment to the subdivision F guideline requirements to provide for an evaluation of the immunotoxicity of chemical pesticides

proposals is a flow diagram for the evaluation of results from this basic study for hazard identification and conclusions drawn from it. This flow diagram, shown in Figure 15.1, could be helpful for the evaluation of results obtained from incorporation of a basic immunotoxicity test battery into studies with routinely treated animals. References BASKETTER, D. et al., 1994, The identification with sensitising or immunosuppressive properties in routine toxicology, Food Chem. Toxicol. (in press). BOTHAM, P.A. et al., 1991, Skin sensitization—a critical review of predictive test methods in animals and men Food Chem. Toxicol, 29, 275–86. BÜHLER, E.V., 1965, Delayed contact hypersensitivity in the guinea pig, Arch. Dermatol. 91, 171. DEAN, J.H., PADARATHSINGH, M.L. and JERRELS, T.R., 1979, Assessment of immunobiological effects induced by chemicals, drugs or food additives. I. Tier testing and screening approach, Drug Chem. Toxicol. 2, 5–17. ECETOC, 1990, Skin Sensitization Testing, Monograph No. 14, Brussels.

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Figure 15.1 Flow diagram (taken from the ECETOC monograph: Immunotoxicity). ECETOC, 1994, Immunotoxicity: hazard identification and risk assessment, Monograph No. 21, Brussels. EEC, 1991, Annex to directive 81/852/EEC—Note of guidance, 22.

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HINTON, D.M., 1992, Testing guidelines for evaluation of the immunotoxic potential of direct food additives, Crit. Rev. Food Sci. Nutrit, 32, 173–90. KIMBER, I., HILTON, J. and WEISENBERGER, C. 1989, The murine local lymph node assay for identification of contact allergens: a preliminary evaluation of in situ measurement of lymphocyte proliferation, Contact Dermatit. 21, 215– 20. LUSTER, M.I. et al., 1988, Development of a battery to assess chemical-induced immunotoxicity: National Toxicology Program’s guidelines for immunotoxicity evaluation in mice, Fundam. Appl. Toxicol. 10, 2–19. LUSTER, M.I. et al., 1992, Risk assessment in immuno-toxicology. I. Sensitivity and predictability of immune tests, Fundam. Appl. Toxicol., 18, 200–10. LUSTER, M.I. et al., 1993, Risk assessment in immuno-toxicology. II. Relationship between immune and host resistance tests, Fundam. Appl. Toxicol., 21, 71–82. MAGNUSSON, B. and KLIGMANN, A.M., 1969, The identification of contact allergens by animal assay. The guinea pig maximisation test, J. Invest. Dermatol., 52, 268. OECD, 1992a, Organisation for Economic Cooperation and Development, Guidelines for testing of chemicals No. 407, adopted 12 July, 1992. OECD, 1992b, Organisation for Economic Cooperation and Development, Guidelines for testing of chemicals—skin sensitisation, No. 406, adopted 17 July, 1992. SJOBLAD, R., 1988, Potential future requirements for immunotoxicology testing of pesticides, Toxicol. Indust. Hlth, 4, 391. TRIZIO, D. et al., 1988, Identification of immunotoxic effects of chemicals and assessment of their relevance to man. Food Chem. Toxic., 26, 527–39. UK Department of Health, 1991, Proposed to update OECD Guideline 407. US-EPA, 1982, Code of Federal Regulations, Washington DC, 152–18, 152–24 and 158–165. US-EPA, 1990, Draft immunotoxicity study screen for testing chemical pesticides. VAN LOVEREN, H. and Vos, J.G., 1992, Evaluation of OECD Guideline 407 for assessment of toxicity of chemicals with respect to potential adverse effects to the immune system. RIVM Report No. 158801001, Bilthoven: National Institute of Public Health and Environmental Protection. VOHR, H.-W., 1995, Experiences with an advanced screening procedure for the identification of chemicals with an immunotoxic potential in routine toxicology (a position paper). Toxicology, (in press),

16 New Strategies: the Use of Long-term Cultures of Hepatocytes in Toxicity Testing and Metabolism Studies of Chemical Products Other than Pharmaceuticals VERA ROGIERS,1 MAY AKRAWI,2 SANDRA COECKE,1 YVES VANDENBERGHE,1 ELIZABETH SHEPHARD,2 IAN PHILLIPS3 and ANTOINE VERCRUYSSE1 1

Vrije Universiteit Brussel, Brussels; 2 University College London, London; 3 University of London, London Introduction: metabolism and toxicity of chemical products are closely linked

Lipophilic compounds are metabolized in the liver by phase 1 and phase 2 reactions into more polar, more hydrophilic metabolites, which are usually less biologically active. Bioactivation, however, may occur, forming toxic species by phase 1, cytochrome P450 (CYP) dependent oxidation (e.g. epoxidation of C=C to reactive epoxide intermediates (Guengerich et al. 1991), CYP dependent reduction (e.g. dehalogenation of CCl4 toa free radical intermediate) (Timbrell, 1993) or even by phase 2 reactions (e.g. reactive episulphonium ion formation by glutathione conjugation of dibromoethane) (Van Bladeren, 1988; Coles and Ketterer, 1990; Timbrell, 1993). The major process involved in the bioactivation of chemical carcinogens is their oxidation catalyzed by CYP enzymes. Thirty or more different CYP forms exist within each animal species (Nebert et al., 1989), each with at least some distinct elements of catalytic specificity and regulation. The role of some of these CYPs in the activation and detoxication of chemical carcinogens has already been determined. For example: – CYP2E1 is a major catalyst involved in the oxidation of benzene, styrene, CCl4, CHCl3, ethylene dichloride, vinylchloride, acrylonitrile, vinyl carbamate (Guengerich et al., 1991), ethanol (Perrot et al., 1989), dialkylnitrosamines (Yoo et al., 1988), isobutene (Cornet et al., 1991) and some other small molecules. – CYP1A1 is involved in the oxidation of polycyclic aromatic hydrocarbons (Shimada et al., 1989b). – CYP1A2 activates arylamines (Shimada et al., 1989a).

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– CYP3A4 is a major catalyst in the activation of aflatoxins, pyrrolizidine alkaloids and polycyclic hydrocarbon dihydrodiols (Shimada et al., 1989a; Shimada and Guengerich, 1989). The balance between the rates of formation of reactive metabolites and detoxication will greatly determine the potential toxic response of a chemical. Variables, known to affect normal biotransformation, such as enzyme induction and inhibition can also change the bioactivation rate of chemicals. A clear example is the metabolism of halogenated biphenyls after treatment with arochlor 1254 (Borlakoglu and Wilkins, 1993). Consequently, the toxicity of chemical products often depends upon their specific biotransformation, the presence, absence, induction and inhibition of specific phase 1 and phase 2 enzymes involved in their metabolism. Towards an in vitro approach of risk assessment Nearly all toxicological studies on chemical products, including industrial chemicals, agrochemicals, pharmaceuticals, additives, materials in contact with food and cosmetics, have been carried out in vivo using experimental animals, in particular small vertebrates (News and Views, 1993). Important scientific, technological, ethical and economic considerations, however, justify the actual search for in vitro alternatives, replacing or improving existing in vivo methods and reducing the number of animals involved (Frazier and Goldberg, 1990; Roberfroid, 1991). Since hepatocytes can be isolated from different species (GuguenGuillouzo et al., 1982; Green et al., 1986; van’t Klooster et al., 1992) including man (Guguen-Guillouzo et al., 1982; Rogiers, 1993), they can represent a powerful tool for short-term risk assessment studies when used as suspensions of freshly isolated cells (up to 3–4 h) or as short-term cultures (up to 2 days) (Klaassen and Stacey, 1982; Guillouzo, 1986). They can be useful in biotransformation, cytotoxicity, hepatotoxicity and genotoxicity studies, in species selection and in mechanistic studies (Blaauboer, 1994). Long-term cultures of hepatocytes (up to several weeks) represent a somewhat different approach in risk assessment. Such systems are of interest for the assessment of long-term toxicity of xenobiotics including the occurrence of enzyme induction, the effects on xenobiotic biotransformation, lipid peroxidation, accumulation of triglycerides, changes in glutathione content, interaction between compounds and the hepatoprotection afforded by certain molecules. Consequently, long-term cultures have been particularly applied in the development of pharmaceuticals, in pharmaco-toxicological studies (Guillouzo, 1986, 1992; Rogiers and Vercruysse, 1993; Skett, 1995). As far as chemical products other than pharmaceuticals and in particular industrial chemicals and agrochemicals are concerned, the practical needs during development

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are different. The compounds brought onto the European market are labelled and need to fulfill only the legal demands of the specific category to which they belong. Testing is only obligatory when they reach the market and the type of tests needed per category of chemicals is clearly outlined. In Europe legal categories consist of dangerous compounds (88/379/EEC, 93/18/EEC),1a,b phytopharmaceuticals (91/414/EEC, 93/71/EEC),2 biocides (93/C239/03),3 cosmetics (76/ 768/EEC, 93/35/5/EEC)4a,b and food additives. Of the latter category the most comprehensive surveys are carried out by the Joint Expert Committee on Food Additives (JECFA) of the World Health Organization and the Food and Agriculture Organization of the United Nations (Conning, 1993). Less sophisticated in vitro studies have been performed on industrial chemicals and agrochemicals than is the case for pharmaceuticals. The only field in which isolated hepatocytes have already been incorporated into routine screening of industrial chemicals for regulatory purposes is in genotoxicity testing (Swierenga et al., 1991). The potentialities, however, of in vitro testing for these compounds, in particular of the use of long-term cultures, has not yet been explored in depth, although induction, inhibition, biotransformation, chronic toxicity, interaction between chemicals and mechanistic studies are of great interest for these compounds too. Human exposure to chemical products such as pesticides, eventually reaching the food chain as residues or of potential risk for workers and operators spraying the fields, is such an interesting research area. For example, Alachlor®, a herbicide, of which millions of tons are used per year, has been classified as a potential human carcinogen (Leslie et al., 1989) because of tumour formation in rats and DNA damage observed in isolated rat hepatocytes (Bonfanti et al., 1992). In vivo studies concerning its biotransformation in rat, mouse and monkey, however, pointed to the observation that Alachlor® is metabolized via different pathways in rodents and monkeys, suggesting a lower risk for man than assumed (internal report Monsanto, 1988). In the future, this type of biotransformation study could easily be performed with short- and long-term cultures of hepatocytes derived from different species, including man, providing relevant human information without interspecies extrapolation. Long-term hepatocyte cultures During culture, hepatocytes undergo phenotypic changes as a function of culture time affecting selectively components of phase 1 and/or phase 2 biotransformation (Nakamura et al., 1983; Guillouzo, 1986; Mooney et al., 1992; Kocarek et al., 1993; Rogiers and Vercruysse, 1993). These changes are interpreted as dedifferentiation. In the literature, data have

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been presented that part of this loss of functionality can be prevented by several factors including soluble medium factors, extracellular matrix components and cell-cell interactions (reviews Guillouzo et al., 1990; Rogiers, 1993). At present no ‘ideal’ long-term hepatocyte culture model exists but valuable alternatives are being developed which take into account some of the factors mentioned. A recent development consists of hepatocytes cultured in a collagen gel sandwich configuration (Dunn et al., 1991; Lee et al., 1992). This promising system is claimed to maintain longterm differentiation probably due to the reinstatement of the cellular polarity of the hepatocytes as a function of the extracellular matrix (Dunn et al., 1991; Lee et al., 1992). To date, only few results concerning xenobiotic metabolism, are available (Koebe et al., 1994). Another recently introduced model substantially improved the maintenance of xenobiotic metabolism by culturing hepatocytes on a mixture of crude membrane fractions with collagen type 1, combined with the use of culture medium supplemented with aprotinin and selenium (Saad et al., 1993). Also co-cultures of hepatocytes with rat epithelial cells of primitive biliary origin represent a rather new and valuable tool in xenobiotic biotransformation research and testing (Bégué et al., 1984a). The model has been developed in order to mimic better the microenvironment of the liver cells in vivo. It is until now, the only long-term hepatocyte culture system of which enough biotransformation data exist. Co-cultures of hepatocytes retain to a great extent the morphological and biochemical characteristics of adult hepatocytes in vivo, including phase 1 and phase 2 xenobiotic metabolism pathways (Guillouzo, 1986; Rogiers et al., 1992; Akrawi et al., 1993a). The following text reviews the actual knowledge concerning xenobiotic biotransformation in cocultured hepatocytes with emphasis on the authors’ own research. Phase 1 reactions in co-cultured hepatocytes It has been claimed that as much as 100 per cent of the CYP content and Naminopyrine demethylation activity can be maintained in co-cultured rat hepatocytes with primitive biliary duct cells (Bégué et al., 1984b). Some other phase 1 enzymatic activities also appear to be maintained since drugs such as ketotifen (Le Bigot et al., 1987) and testosterone (Utesch, 1992) were metabolized by several pathways. Maier (1988) however, has reported that aldrin epoxidase activity underwent a significant decrease as a function of culture time. Niemann et al. (1991) was able to show that, as was also the case for mono-cultured rat hepatocytes, enzymatic activities belonging to the 3-methyl-cholanthrene-inducible family were better maintained than those belonging to the phenobarbital inducible family. In our own experiments with adult rat hepatocytes co-cultured with rat liver

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epithelial cells (Rogiers et al., 1990b), it was found that a steady-state situation for at least 10 days was obtained in which the total CYP content and 7-ethoxycoumarin O-deethylase and aldrin epoxidase activities were maintained at 25,100 and 15 per cent, respectively, of their corresponding values in freshly isolated hepatocytes. Both the total CYP content and 7ethoxycoumarin O-deethylase activity, but not the aldrin epoxidase activity, were induced by exposure to phenobarbital (Rogiers et al., 1990b) or sodium valproate (Rogiers et al., 1988b). Furthermore, the combination of inducing agents with co-cultivation of rat hepatocytes with rat epithelial cell clones has recently been proposed as the best way for stabilization of the CYP system. This method is better than the use of a perfusion system or changing the extracellular matrix from collagen to matrigel (Wegner et al., 1991). The results mentioned above indicate a degree of selection in the ability of co-cultured hepatocytes to maintain the expression and inducibility of individual members of the CYP superfamily. From the work of Akrawi (Akrawi et al., 1993a), using mRNA analysis of co-cultured rat hepatocytes, it appeared that the abundance of CYP2B mRNAs declined to about 30 per cent of the initial value by 4 days but that thereafter it remained constant. The inducibility by phenobarbital (Akrawi et al., 1993a) and sodium valproate (Rogiers et al., 1992) was also maintained. These results were confirmed by Western blotting (Akrawi et al. 1993b). RNase protection assays using probes capable of distinguishing between CYP2B1 and CYP2B2 mRNAs demonstrated that the relative abundance and inducibility of each of the mRNAs were the same in co-cultures as in vivo. Co-cultured hepatocytes also maintained the expression of the CYP1A2 gene and of genes coding for two other components of the CYP-mediated monooxygenase, namely NADPH cytochrome P450 reductase and cytochrome b5 (Akrawi et al., 1993a). Both components were inducible by valproate and phenobarbital (Akrawi et al., 1993b; Shephard et al., 1994; Rogiers et al., 1994). In addition, we have shown using Western blotting (Akrawi et al., 1993b) and mRNA analysis (Akrawi et al., 1994; Shephard et al., 1994) that the expression of CYP4A and its specific inducibility (induced by valproate and not by phenobarbital) were maintained in co-cultured rat hepatocytes. By using antisense RNA probes that could discriminate between RNAs encoding different members of the CYP4A subfamily it was further demonstrated that CYP4A1, CYP4A2 and CYP4A3 were all induced by valproate, although to differing extents. None of these mRNAs was increased by phenobarbital (Akrawi et al., 1994; Shephard et al. 1994). The results were very similar to those observed in vivo (Shephard et al., 1994). Flavin-containing monooxygenase (FMO), a less well-known phase 1 biotransformation enzyme than the CYP system, is responsible for the oxygenation of drugs, pesticides and dietary components (Ziegler, 1980). It

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may activate as well as deactivate a number of important molecules such as thioether-containing pesticides (Hajjar and Hodgson, 1980), 3,3′-dichlorobenzidine (Iba and Thomas, 1988), N-methyl-4-aminobenzene (Kadlubar et al., 1976) and others. The expression of FMOs is also maintained better and for a longer time in co-cultures of rat hepatocytes than it is in monocultures of these cells (Coecke et al., 1993). In co-cultures a steady state situation is obtained at a level of approximately 40 per cent of its initial value in freshly isolated hepatocytes (Coecke et al., 1993). Hormonal regulation of FMO is retained (Coecke et al., 1995a,b). It was shown that 17β-oestradiol significantly decreased FMO activity in cocultures of male rat hepatocytes which was not the case for testosterone and 5 α-dihydrotestosterone. These data are in accordance with in vivo results (Coecke et al., 1995b). In addition, the thyroid hormone thyroxine and its metabolite L-triiodothyronine were found to cause a significant decrease in FMO activity suggesting a suppressive role in the regulation of FMO in rat liver. In vivo data on this subject are not available in the literature. Phase 2 reactions in co-cultured hepatocytes The activity, expression and regulation of the different glutathione Stransferase (GST) isoenzymes in co-cultured rat hepatocytes, has been extensively studied by our group (Vandenberghe et al., 1988a,b, 1989, 1990a,b, 1991). As far as the enzymatic activity is concerned, it was observed that the composition of the culture medium was of much less influence in co-cultures than was the case in mono-cultures (Vandenberghe et al., 1988b). In cocultures GST activity was maintained for a longer period and at a more stable level, comparable to the in vivo situation. This observation was confirmed later by Niemann et al. (1991) and Utesch and Oesch (1992), although the latter investigators reported a high variability in their results depending on the batch of epithelial cells. GST activity in cocultured rat hepatocytes was found to be increased or decreased by phenobarbital and valproate, respectively (Rogiers et al., 1988a; Rogiers et al., 1992). These results are in good accordance with previously obtained in vivo data (Rogiers et al., 1988a; Rogiers et al., 1992). They were also confirmed at the protein level using the Western blot technique (Rogiers et al., 1995). Furthermore, GST activity is increased significantly, in a dose dependent way by ethanol. Both GST protein and mRNA amounts (in particular, GST subunits 3 and 4) were increased by this compound (Coecke et al., 1995c). Results obtained using different substrates suggested that the GST isoenzyme profile changes as soon as hepatocytes are seeded in culture (Vandenberghe et al., 1988b). By a combination of GSH agarose affinity

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chromatography and reversed phase HPLC, the GST subunits of cocultured rat hepatocytes were purified and separated (Vandenberghe et al, 1988a). Alterations comparable to those observed for mono-cultures suggested changes towards a more ‘foetal-like’ state. Less variations, however, were noticed in the GST subunit pattern of co-cultured hepatocytes when various media conditions were compared. Incorporation of 35S-methionine in the medium showed the ability of co-cultured rat hepatocytes to synthesize the different GST subunits and suggested that changes in GST subunit expression under various culture conditions were the result of in vitro ‘de novo’ synthesis (Vandenberghe et al., 1990a). Northern blot analysis, using specific cDNA probes showed that the mRNA levels encoding GST subunits 1/2, 3/4 and 7 were very dependent on the culture medium. Again in co-cultures, the changes observed were much less marked than was the case for mono-cultures (Morel et al., 1989; Vandenberghe et al., 1990b). As already mentioned for conventionally cultured rat hepatocytes, phenobarbital had inducing effects on all the GST subunits, but to a different extent for each subunit (Vandenberghe, 1989). The increased steady-state mRNA levels observed in co-cultures after phenobarbital exposure were the result of an increased transcriptional activity of the GST genes together with a stabilizing effect of the compound (Vandenberghe et al., 1991). Also of interest is that the hormonal regulation of GST is maintained in co-cultures of male rat hepatocytes. 17β-Oestradiol, triiodothyronine and thyroxine cause a significant decrease in GST activity. Both the overall GST activity and in particular that of GST 3–3 and 3–4 are decreased (Coecke et al., 1995c). In contrast, male sex hormones and human growth hormone had little effect on the overall activity. The effects of triiodothyronine and thyroxine were particularly oriented towards GST subunits 3 and 4 and towards an as yet unidentified GST subunit, which was significantly increased (Coecke et al., 1995c). 17β-Oestradiol shifted the GST subunit pattern towards the one observed in freshly isolated cells whereas growth hormone had no specific effect on the individual protein classes (Coecke et al., 1995c). These results clearly show a hormonal regulation of GST in co-cultured rat hepatocytes, although previous work with mono-cultures failed to prove any direct effect (Gebhardt et al., 1990). The effects are most pronounced for the Mu-class GSTs. In man, Mu-class GST genes are structurally very similar to the rat genes and are of particular interest because 45 per cent of the European population fails to express a transferase at the GST M1 locus (Zhong et al., 1993). It is this class of GSTs that is very effective in deactivating mutagenic and carcinogenic epoxides. UDP glucuronyltransferases (UDP-GT) have been much less studied in cocultures than GST. From the work of Niemann et al. (1991) it appears

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that 1-naphthol UDP-GT activity is maintained well in co-cultures whereas this is not true for morphine UDP-GT, although for both a steady state situation was reached. These data point to a shift towards a more differentiated pattern since 1-naphthol is considered to be more specific for the late foetal form of UDP-GT and morphine for the neonatal form. The maintenance of 1-naphthol UDP-GT has been confirmed by Utesch and Oesch (1992). Other studies on preservation of phase 2 enzymatic activity in co-cultures have dealt with the identification of the metabolites formed when drugs are added to the culture medium. In human co-culture it could be shown that the glucuronide metabolite of ketotifen was still present after 3 weeks whereas it became undetectable after 6 days in mono-culture (Bégué et al., 1984b). Similar observations have been made for other drugs such as caffeine and theophylline (Ratanasavanh et al., 1990). Conclusions At present no ideal culture system for hepatocytes can be proposed. In all models reported in the literature, phenotypic changes occur, affecting the various components of phase 1 and phase 2 xenobiotic metabolism to a different extent. An interesting conclusion, however, remains from the observation that, when co-cultures and mono-cultures of hepatocytes are compared, cocultures exhibit higher biotransformation capacities which are better and preserved for longer than is the case for mono-cultures. The inducibility by common inducers is fairly well maintained and seems, to a certain extent, comparable with the in vivo situation. In addition, hormonal regulation of phase 1 and phase 2 key enzymes seems to be well maintained and comparable with the in vivo situation. Co-cultures of hepatocytes with rat liver epithelial cells are therefore already of importance as an alternative model for risk assessment. In particular, when long-term effects of a chemical are to be expected. Some experience already exists concerning the application of co-cultured hepatocytes for the study of pharmaceuticals. As far as chemical products other than pharmaceuticals are concerned, experience is lacking although interesting results are to be expected particularly in those cases where chemicals can interfere with the human organism via the food chain or by occupational exposure. In vitro exploration of this new field in toxicology is a challenge for the coming years. Notes 1a. 88/379/EEC Directive du Conseil du 7 juin 1988 concernant le rapprochement des dispositions législatives, réglementaires et administratives des Etats membres

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1b.

2a.

2b.

3.

4a.

4b.

relatives à la classification, à l’emballage et à 1’étiquetage des préparations dangereuses. Journal Officiel des Communautés européennes no L187, 16 juillet 1988, p. 14. 93/18/EEC Directive 93/18/CEE de la Commission du 5 avril 1993 portant troisième adaptation au progrès technique de la directive 88/379/CEE du Conseil concernant le rapprochement des dispositions législatives, réglementaires et administratives des Etats membres relatives à la classification, à l’emballage et à 1'étiquetage des preparations dangereuses. Journal Officiel des Communautés européennes no L104, 29 Avril 1993, p. 46. 91/414/EEC Richtlijn van de Raad van 15 juli 1991 betreffende het op de markt brengen van gewasbeschermingsmiddelen. Publikatieblad van de Europese Gemeenschappen no L230, 19 Augustus 1991, p. 1. 93/71/EEC Directive 93/71/CEE de la Commission du 27 juillet 1993 modifiant la directive 91/414/CEE du Conseil concernant la mise sur le marché des produits phytopharmaceutiques. Journal Officiel des Communautés européennes no L221, 31 Août 1993, p. 27. 93/C239/03 Voorstel voor een richtlijn van de Raad betreffende het op de markt brengen van biociden. Publicatieblad van de Europese Gemeenschappen no C239, 3 September 1993, p. 3. 76/768/EEC Richtlijn van de Raad van 27 juli 1976 betreffende de onderlinge aanpassing van de wetgevingen der Lid-Staten inzake kosmetische produkten. Publikatieblad van de Europese Gemeenschappen no L262, 27 juli 1976, p. 169. 93/35/EEC Directive 93/35/CEE du Conseil du 14 juin 1993 modifiant pour la sixième fois la directive 76/768/CEE concernant le rapprochement des lé-gislations des Etats membres relatives aux produits cosmétiques. Journal Officiel des Communautés européennes no L151, 23 juin 1993 p. 32.

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NIEMANN, C., GAUTHIER, J.C., RICHERT, L., IVANOV, M.A., MELCION, C. and CORDIER, A., 1991, Rat adult hepatocytes in primary pure and mixed monolayer culture, Biochemical Pharmacology, 42, 373–9. PERROT, N., NALPAS, B., YANG, C.S. and BEAUNE, P.H., 1989, Modulation of cytochrome-P450 isozymes in human liver, by ethanol and drug intake, European Journal of Clinical Investigation, 19, 549–55. RATANASAVANH, D., BAFFET, G., LATINIER, M.F., RISSEL, M. and GUILLOUZO, A., 1988, Use of hepatocyte co-cultures in the assessment of drug toxicity from chronic exposure, Xenobiotica, 18, 765–71. RATANASAVANH, D., BERTHOU, F., DREANO, Y., MONDINE, P., GUILLOUZO, A. and RICHE, C., 1990, Methylcholanthrene but not phenobarbital enhances caffeine and theophylline metabolism in cultured human hepatocytes, Biochemical Pharmacology, 266, 683–8. ROBERFROID, M.B., 1991, Long term policy in toxicology, in Hendriksen, C.F.M. and Koëter, H.B.W.M. (Eds) Animals in Biochemical Research, pp. 35–48, Amsterdam: Elsevier Science Publishers B.V. ROGIERS, V., 1993, Cultures of human hepatocytes in in vitro pharmacotoxicology, in Rogiers, V., Sonck, W., Shephard, E. and Vercruysse, A. (Eds) Human Cells in in vitro Pharmaco-toxicology—Present Status within Europe, pp. 77–115, Brussels: VUB Press. ROGIERS, V., VANDENBERGHE, Y., CALLAERTS, A., SONCK, W., MAES, V. and VERCRUYSSE, A., 1988a, The inducing and inhibiting effects of sodium valproate in vivo on the biotransformation systems of xenobiotics in isolated rat hepatocytes, Xenobiotica, 18, 665–73. ROGIERS, V., VANDENBERGHE, Y., CORNET, M., CALLAERTS, A., SONCK, W., GUILLOUZO, A. and VERCRUYSSE, A., 1988b, The use of cultures and cocultures of rat hepatocytes for the study of the effects of valproate on phase I and II biotransformation systems of xenobiotics, in Woodhouse, K.W., Yelland, C. and James, O.F.W. (Eds) The Liver, Metabolism and Ageing, pp. 75–82, Amsterdam: Elsevier Biomedical Press. ROGIERS, V., VANDENBERGHE, Y., CALLAERTS, A., SONCK, W. and VERCRUYSSE, A., 1990a, Effects of dimethylsulphoxide on phase I and phase II biotransformation in cultured rat hepatocytes, Toxicology In Vitro, 4, 239– 42. ROGIERS, V., VANDENBERGHE, Y., CALLAERTS, A., VERLEYE, G., CORNET, M., MERTENS, K., SONCK, W. and VERCRUYSSE, A., 1990b, Phase I and phase II xenobiotic biotransformation in cultures and co-cultures of adult rat hepatocytes, Biochemical Pharmacology, 40, 1701–6. ROGIERS, V., CALLAERTS, A., VERCRUYSSE, A., AKRAWI, M., SHEPHARD, E. and PHILLIPS, I., 1992, Effects of valproate on xenobiotic biotransformation systems in rat liver: in vivo and in vitro experiments, Pharmaceutisch Weekblad, Scientific Edition, 14, 117–31. ROGIERS, V. and VERCRUYSSE, A., 1993, Rat hepatocyte cultures and cocultures in biotransformation studies of xenobiotics, Toxicology, 82, 193–208. ROGIERS, V., AKRAWI, M., VERCRUYSSE, A., SHEPHARD, E. and PHILLIPS, I. R., 1995, Effects of valproate on the expression of the cytochrome P450 system in rat liver and co-cultured rat hepatocytes, Biochemical Journal, in press.

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SAAD, B., SCHOLL, F.A., THOMAS, H., SCHAWALDER, H., STREIT, V., WAECHTER, F. and MAIER, P., 1993, Crude liver membrane fractions and extracellular matrix components as substrata regulate differentially the preservation and inducibility of cytochrome P-450 isoenzymes in cultured rat hepatocytes, European Journal of Biochemistry, 213, 805–14. SHEPHARD, E.A., AKRAWI, M., ROGIERS, V., BOOTMAN, M., VERCRUYSSE, A. and PHILLIPS, I.R., 1995, Valproate increases the expression of the CYP4A subfamily in vivo and in vitro, Biochemical Journal, submitted. SHIMADA, T. and GUENGERICH, F.P., 1989, Evidence for cytochrome P-450 NF, the nifedipine oxidase, being the principal enzyme involved in the bioactivation of aflatoxins in human liver, Proceedings of the National Academy of Sciences of the United States of America, 86, 462–5. SHIMADA, T., IWASAKI, M., MARTIN, M.V. and GUENGERICH, F.P., 1989a, Human liver microsomal cytochrome P-450 enzymes involved in the bioactivation of procarcinogens detected by umu gene response in Salmonella typhimurium TA1535/pSK 1002, Cancer Research, 49, 3218–28. SHIMADA, T., MARTIN, M.V., PRUESS-SCHWARTZ, D., MARNETT, L.J. and GUENGERICH, F.P., 1989b, Roles of individual human cytochrome P-450 enzymes in the bioactivation of benzo(a)pyrene, 7,8-dihydroxy-7,8dihydrobenzo(a) pyrene, and other dihydrodiol derivatives of polycyclic aromatic hydrocarbons, Cancer Research, 49, 6304–12. SKETT, P., 1995, Problems in using isolated and cultured hepatocytes for xenobiotic metabolism/metabolism-based toxicity testing—Solutions? Toxicology in Vitro (in press). SWIERENGA, S.H.H., BRADLAW, J.A., BRILLINGER, R.L., GILMAN, J.P. W., NESTMANN, E.R. and SAN, R.C., 1991, Recommended protocols based on a survey of current practice in genotoxicity testing laboratories: I. Unscheduled DNA synthesis assay in rat hepatocyte cultures, Mutation Research, 246, 235– 53. TEE, L.B.G., DAVIES, D.S., SEDDON, C.E. and BOOBIS, A.R., 1987, Species differences in the hepatotoxicity of paracetamol are due to differences in the rate of conversion to its cytotoxic metabolite, Biochemical Pharmacology, 36, 1041–52. TIMBRELL, J., 1993, Biotransformation of xenobiotics, in Ballantyne, B.(Marrs, T. and Turner, P. (Eds) General and Applied Toxicology, pp. 88–119, New York: Stockton Press. UTESCH, D. and OESCH, F., 1992, Dependency of the in vitro stabilization of differentiated functions in liver parenchymal cells on the type of cell line used for coculture, In vitro Cellular Development Biology, 28, 193–8. VAN BLADEREN, P., 1988, Formation of toxic metabolites from drugs and other xenobiotics by glutathione conjugation, Trends in Pharmacological Sciences, 9, 295–8. VANDENBERGHE, Y., GLAISE, D., MEYER, D., GUILLOUZO, A. and KETTERER, B., 1988a, Glutathione transferase isoenzymes in cultured rat hepatocytes, Biochemical Pharmacology, 37, 2482–95. VANDENBERGHE, Y., RATANASAVANH, D., GLAISE, D. and GUILLOUZO, A., 1988b, Influence of medium composition and culture conditions on

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glutathione S-transferase activity in adult rat hepatocytes during culture, In Vitro Cellular and Developmental Biology, 24, 281–8. VANDENBERGHE, Y., MOREL, F., FORIERS, A., KETTERER, B., VERCRUYSSE, A., GUILLOUZO, A. and ROGIERS, V., 1989, Effect of phenobarbital on the expression of glutathione S-transferase isoenzymes in cultured rat hepatocytes, FEBS Letters, 251, 59–64. VANDENBERGHE, Y., FORIERS, A., ROGIERS, V. and VERCRUYSSE, A., 1990a, Changes in expression and ‘de novo’ synthesis of glutathione Stransferase subunits in cultured adult rat hepatocytes, Biochemical Pharmacology, 39, 685–90. VANDENBERGHE, Y., MOREL, F., PEMBLE, S., TAYLOR, J.B., ROGIERS, V., RATANASAVANH, D., VERCRUYSSE, A., KETTERER, B. and GUILLOUZO, A., 1990b, Changes in expression of mRNA coding for glutathione S-transferase subunits 1–2 and 7 in cultured rat hepatocytes, Molecular Pharmacology, 37, 372–6. VANDENBERGHE, Y., TEE, L., MOREL, F., ROGIERS, V., GUILLOUZO, A. and YEOH, G., 1991, Regulation of glutathione S-transferase gene expression by phenobarbital in cultured adult rat hepatocytes, FEBS Letters, 284, 103–8. VAN’T KLOOSTER, G.A.E., WOUTERSEN-VAN NIJANTEN, F.M.A., KLEIN, W.R., BLAAUBOER, B.J., NOORDHOEK, J. and VAN MIERT, A.S.J.P.A. M., 1992, Effects of various medium formulations and attachment substrata on the performance of cultured ruminant hepatocytes in biotransformation studies, Xenobiotica, 22, 523–34. WEGNER, H., MECKE, D. and GEBHARDT, R., 1991, Attempts to stabilize the cytochrome P450 dependent monooxygenase activity in primary cell-culture. Abstract in Third International ISSX meeting: Drug metabolism: molecules, models and man, p. 317, Amsterdam, June. Yoo, J.S.H., GUENGERICH, F.G. and YANG, C.S., 1988, Metabolism of Nnitrosodialkylamines by human liver microsomes, Cancer Research, 48, 1499– 504. ZHONG, S., SPURR, N.K., HAYES, J.D. and WOLF, C.R., 1993, Deduced amino acid sequence gene structure and chromosome location of a novel human class Mu glutathione S-transferase, GSTM4, Biochemical Journal, 291, 41–50. ZIEGLER, D.M., 1980, Microsomal flavin-containing monooxygenase: oxygenation of nucleophilic nitrogen and sulfur compounds, in Jakoby, W.B. (Ed.) Enzymatic Basis of Detoxification, pp. 201–7, New York: Academic Press.

PART FIVE Mechanisms of toxicity of industrial chemicals

17 Peroxisome Proliferation BRIAN G.LAKE and ROGER J.PRICE BIBRA International, Carshalton, Surrey

Introduction Peroxisomes (sometimes referred to as ‘microbodies’) are single membranelimited cytoplasmic organelles which are characterised by their content of catalase and a number of hydrogen peroxide generating oxidase enzymes (Cohen and Grasso, 1981; Reddy and Lalwani, 1983). In rat liver, peroxisomes are normally spherical or oval in shape, approximately 0.5 µm in diameter and contain a finely granular matrix with a crystalline nucleoid core. A number of reviews have been published dealing with various aspects of hepatic peroxisome proliferation (Cohen and Grasso, 1981; Reddy and Lalwani, 1983; Hawkins et al., 1987; Stott, 1988; Lock et al., 1989; Moody et al., 1991; Bentley et al., 1993; Lake, 1993). This chapter will focus on mechanisms of hepatocarcinogenesis, species differences in response and risk assessment of rodent peroxisome proliferators. Peroxisome proliferation in rodent liver Since the initial observations on the hepatic effects of the hypolipidaemic agent clofibrate (Paget, 1963; Hess et al., 1965) many compounds have been shown to produce hepatic peroxisome proliferation in rats and mice. Liver enlargement is due to both hyperplasia and hypertrophy and organelle proliferation is associated with a differential induction of peroxisomal enzyme activities. Peroxisomes, like mitochondria, contain a complete fatty acid β-oxidation cycle (Lazarow and DeDuve, 1976). While the enzymes of the β-oxidation cycle (normally assessed as cyanideinsensitive palmitoyl-CoA oxidation) are markedly induced, only small changes are observed in other peroxisomal enzyme activities such as catalase and D-amino acid oxidase. Apart from stimulating peroxisomal fatty acid metabolism, peroxisome proliferators also increase microsomal fatty acid ( -l)- and particularly -hydroxylase activities. This is due to induction of cytochrome P-450 isoenzymes in the CYP4A subfamily and is

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normally measured as lauric acid 12-hydroxylase (Sharma et al., 1988a, b; Gibson, 1989). Peroxisome proliferators also markedly increase carnitine acetyltransferase activity which is localised in peroxisomal, mitochondrial and microsomal fractions (Ishii et al., 1980; Bieber et al., 1981). In rat liver good correlations have been reported between the induction of peroxisomal fatty acid β-oxidation and organelle proliferation and between the induction of peroxisomal and microsomal fatty acid oxidising enzyme activities (Lake et al., 1984a; Lin, 1987; Sharma et al., 1988a, b; Dirven et al., 1992). Several laboratories have demonstrated that the characteristics of peroxisome proliferation in vivo may also be observed in vitro in primary rat and mouse hepatocyte cultures. Indeed, hepatocyte cultures have been employed for studying various aspects of peroxisome proliferation including structureactivity relationships and species differences in response (Gray et al., 1982; Elcombe, 1985; Bieri, 1993; Lake and Lewis, 1993; Foxworthy and Eacho, 1994). Rodent peroxisome proliferators Many different classes of chemicals have been found to produce peroxisome proliferation in the rat and mouse (Cohen and Grasso, 1981; Reddy and Lalwani, 1983; Stott, 1988; Moody et al., 1991; Bentley et al., 1993; Lake and Lewis, 1993). Classes of industrial chemicals include plasticisers, chlorinated solvents (e.g. trichloroethylene, perchloroethylene), chlorinated paraffins and other chemicals (e.g. perfluoro-n-octanoic acid). Types of plasticisers known to produce peroxisome proliferation include phthalate esters (e.g. di-(2-ethyl-hexyl)phthalate (DEHP), di-(isodecyl) phthalate), adipate esters (e.g. di-(2-ethyl-hexyl)adipate (DEHA)) and other compounds (e.g. tri-(2-ethylhexyl) trimellitate). Apart from industrial chemicals other known rodent hepatic peroxisome proliferators include herbicides, hypolipidaemic and other categories of therapeutic agents, certain steroids, food flavours and natural products. While peroxisome proliferators appear to be structurally diverse, at least for some compounds, similarities in their three-dimensional structures have been reported (Lake et al., 1988; Lake and Lewis, 1993). Many studies have demonstrated structure-activity relationships for various classes of peroxisome proliferators including industrial chemicals (Lake and Lewis, 1993). A characteristic feature of many, but not all, peroxisome proliferators is the presence of an acidic function (Lake et al., 1988; Lock et al., 1989). This acidic function is normally a carboxyl group, either present as a free carboxyl group in the parent structure or one that is unmasked by metabolism. Alternatively, the chemical may contain a chemical grouping which is a bioisostere of a carboxyl group (Thornber,

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1979), such as tetrazole or a sulphonamide moiety (Eacho et al., 1986; Lock et al., 1989). It should be noted that rodent liver peroxisome proliferators exhibit marked compound potency differences. While potent peroxisome proliferators include compounds developed as hypolipidaemic agents (e.g. ciprofibrate, Wy-14,643), plasticisers such as DEHP are less potent and chemicals such as acetylsalicylic acid are even less potent (Reddy et al., 1986; Barber et al., 1987; Lake and Lewis, 1993). For example, in a 30day feeding study, a similar magnitude of induction of palmitoyl-CoA oxidation was observed in rats fed 0.001 per cent ciprofibrate and 0.5 per cent DEHP diets, whereas for DEHA a dietary level of >1.0 per cent but <2. 0 per cent was required to produce a similar effect (Reddy et al., 1986). Carcinogenicity of peroxisome proliferators Hepatic peroxisome proliferation is of importance, not only because of the large range of industrial and other chemicals which produce this effect in rodents, but because certain of these agents have been found to increase the incidence of liver tumours in rats and/or mice (Cohen and Grasso, 1981; Reddy and Lalwani, 1983; Moody et al., 1991; Bentley et al., 1993). For example, the industrial plasticiser DEHP has been reported to produce liver tumours in rats and mice whereas another plasticiser DEHA and the industrial solvent trichloroethylene have been reported to be hepatocarcinogenic in the mouse (NCI, 1976; NTP, 1982a, b). In addition, peroxisome proliferators may also produce tumours in other organs such as the pancreas and testis (Hinton and Price, 1993). Although peroxisome proliferators can produce hepatocellular carcinoma in rodents, they are not considered to be genotoxic agents. Thus studies with plasticisers and other peroxisome proliferators have shown that they do not bind covalently to DNA after in vivo administration to rats and mice (Von Däniken et al., 1981, 1984; Albro et al., 1983; Goel et al., 1985). Moreover, the short term administration of peroxisome proliferators to rats failed to result in DNA adducts using the sensitive 32Ppostlabelling technique (Gupta et al., 1985). Phthalate esters and other peroxisome proliferators have been extensively tested in a wide range of short-term tests for genotoxicity (Budroe and Williams, 1993). Generally peroxisome proliferators produce negative results in such tests, although a few positive findings have been reported (Budroe and Williams, 1993; Hwang et al., 1993). In keeping with the properties of non-genotoxic rodent hepatocarcinogens, peroxisome proliferators do not produce tumours when examined in initiation studies (Popp and Cattley, 1993). However, when appropriate histochemical markers are employed, several peroxisome proliferators have been

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demonstrated to be effective in rat liver tumour promotion studies (Cattley and Popp, 1989; Bentley et al., 1993; Popp and Cattley, 1993). Mechanisms of hepatocarcinogenesis Several hypotheses have been proposed to account for why peroxisome proliferators can produce liver tumours in rodents. These mechanisms include: (a) Induction of sustained oxidative stress to hepatocytes (Reddy and Lalwani, 1983; Reddy and Rao, 1989). (b) A role of increased cell proliferation (Marsman et al., 1988; Popp and Marsman, 1991). (c) The promotion of spontaneously formed preneoplastic liver lesions (Schulte-Hermann et al., 1989; Cattley et al., 1991; Grasl-Kraupp et al., 1993). (d) A combination of two or all of the above factors. The oxidative stress hypothesis is based on the observation that the chronic administration of peroxisome proliferators produces a sustained oxidative stress in rodent hepatocytes due to an imbalance in the production and degradation of hydrogen peroxide (Reddy and Lalwani, 1983; Reddy and Rao, 1989). Peroxisome proliferators markedly induce the peroxisomal fatty acid β-oxidation cycle, but produce only a small increase in catalase activity. The first enzyme of the β-oxidation cycle, acyl-CoA oxidase, produces hydrogen peroxide and hence the cyclic oxidation of a single fatty acid molecule can result in the production of several molecules of hydrogen peroxide (Lazarow and DeDuve, 1976). Any excess hydrogen peroxide not destroyed by peroxisomal catalase can diffuse through the peroxisomal membrane into the cytosol where it will be a substrate for cytosolic selenium-dependent glutathione peroxidase. However, this enzyme activity and that of other enzymes including superoxide dismutase and glutathione S-transferases are often reduced by the administration of peroxisome proliferators to rodents (Reddy and Rao, 1989; Bentley et al., 1993; Lake, 1993). These enzyme changes are postulated to result in increased intracellular levels of hydrogen peroxide which, either directly or via reactive oxygen species (e.g. hydroxyl radical), can attack membranes and DNA (Reddy and Lalwani, 1983; Reddy and Rao, 1989). A number of experimental observations have provided support for the involvement of oxidative stress in the hepatotoxicity of peroxisome proliferators (Reddy and Rao, 1989; Lake, 1993). For example, peroxisome proliferators have been reported in some studies to increase hepatic lipid peroxidation and lipofuscin deposition, to modulate levels of hepatic antioxidants and to increase levels of 8-hydroxydeoxyguanosine in

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hepatic DNA (Reddy and Rao, 1989; Bentley et al., 1993; Lake, 1993). However, the available data suggest that sustained oxidative stress is unlikely to be solely responsible for per oxisome proliferator-induced hepatocarcinogenesis in rodents. Although evidence of oxidative damage to hepatocytes has been observed in some studies, the magnitude of such effects does not correlate with the potency of the compound to produce tumours. For example, oxygen radical attack on DNA is known to result in a variety of modified DNA bases including 8-hydroxydeoxyguanosine. However, at bioassay dose levels both DEHP and DEHA produce similar increases in hepatic 8-hydroxydeoxyguanosine levels, but only DEHP produced liver tumours in male F344 rats (NTP, 1982a, b; Takagi et al., 1990; Lake, 1993). Many studies have demonstrated that cell proliferation is an important factor in the development of tumours by both genotoxic and nongenotoxic agents (Cohen and Ellwein, 1990, 1991). For example, an enhanced rate of cell replication can increase the frequency of spontaneous lesions and the probability of converting DNA adducts from both endogenous and exogenous sources into mutations before they can be repaired (Cohen and Ellwein, 1990, 1991; Popp and Marsman, 1991). Peroxisome proliferators are known to produce a burst of cell replication in rodent hepatocytes during the first few days of administration (Reddy and Lalwani, 1983; Eacho et al., 1991). In some studies peroxisome proliferators have also been shown to produce a sustained stimulation of replicative DNA synthesis (Lake, 1993). Apart from intrinsic compound potency, dose is an important factor in determining whether a particular compound can produce either a transient or a sustained stimulation of replicative DNA synthesis in rodent hepatocytes. For example, low doses of nafenopin and Wy-14,643 do not produce a sustained stimulation of cell replication, whereas higher doses do produce this effect (Eacho et al., 1991; Price et al., 1992; Wada et al., 1992; Lake et al., 1993). Several studies have demonstrated the presence of numerous foci of putative preneoplastic cells in the livers of untreated old rats and mice (Schulte-Hermann et al., 1983; Grasl-Kraupp et al., 1993). These lesions are considered to represent spontaneously initiated cells as they have similar biological characteristics to those of cells initiated by genotoxic carcinogens (Grasl-Kraupp et al., 1993). The ability of peroxisome proliferators to produce tumours in young compared to old rats has been investigated in studies with nafenopin (Kraupp-Grasl et al., 1991) and Wy-14,643 (Cattley et al., 1991). In both studies more adenomas and carcinomas were produced in old as against young rats.

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Species differences in response Many studies have examined species differences in hepatic peroxisome proliferation (Cohen and Grasso, 1981; Rodricks and Turnbull, 1987; Stott, 1988; Lock et al., 1989; Moody et al., 1991; Bentley et al., 1993). Based on both marker enzyme activities and ultrastructural examination the rat and mouse are clearly responsive species, the Syrian hamster appears to exhibit an intermediate response, whereas in most studies the guinea pig is either nonresponsive or refractory. For example, DEHP readily produces peroxisome proliferation in the rat and mouse, to a lesser extent in the Syrian hamster but not in the guinea pig (Osumi and Hashimoto, 1978; Lake et al., 1984b). Similar results have been obtained with more potent compounds including ciprofibrate, clobuzarit, LY 171883 and nafenopin (Orton et al., 1984; Eacho et al., 1986; Lake et al., 1989; Makowska et al., 1992). When assessing species differences in response a number of factors should be considered. These include the metabolism, disposition and dose of the test compound, sex differences, as well as intrahepatic differences in response. The importance of metabolism is illustrated by the industrial solvent trichloroethylene which produces peroxisome proliferation and liver tumours in the mouse but not in the rat (NCI, 1976; Elcombe, 1985). Metabolic studies demonstrated that the trichloroethylene was extensively metabolised to trichloroacetic acid in the mouse, whereas this was a minor saturable route of metabolism in the rat. That the difference in trichloroacetic acid formation was responsible for the observed species difference was demonstrated by the fact that this compound produced peroxisome proliferation in rat and mouse hepatocytes both in vivo and in vitro (Elcombe, 1985). An example of compound disposition is provided by DEHP which is known to be more extensively absorbed after oral administration in the rat than in the marmoset (Rhodes et al., 1986). However, the observed in vivo species differences in response are supported by the observation that metabolites of DEHP which produce peroxisome proliferation in rat hepatocytes in vitro have no significant effect in cultured marmoset hepatocytes (Elcombe and Mitchell, 1986). Generally, in vitro studies with primary hepatocyte cultures from the rat, mouse, Syrian hamster, guinea pig and marmoset have supported the results of in vivo studies in these species (Elcombe 1985; Elcombe and Mitchell, 1986; Lake et al., 1986; Bieri, 1993; Bentley et al., 1993; Foxworthy and Eacho, 1994). Several studies have examined the ability of rodent peroxisome proliferators to produce effects in primates and humans. With respect to primates, studies with a number of compounds in both New (e.g. marmoset) and Old (e.g. Rhesus monkey) World monkeys have failed to provide any evidence of significant hepatic peroxisome proliferation

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(Rodricks and Turnbull, 1987; Bentley et al., 1993). However, albeit at high doses two compounds, namely ciprofibrate (Reddy et al., 1984) and DL-040 (Lalwani et al., 1985), have been reported to produce hepatic peroxisome proliferation in Cynomolgus and/or Rhesus monkeys. In humans, studies have been conducted in patients treated with several hypolipidaemic agents (all being rodent peroxisome proliferators) including ciprofibrate, clofibrate, fenofibrate and gemfibrozil (Bentley et al., 1993). While most studies have failed to detect any significant changes, clofibrate was reported to produce a small increase in the number of peroxisomes (Hanefeld et al., 1983) and ciprofibrate to produce a small increase in the pro portion of the hepatocyte cytoplasm occupied by peroxisomes (cited in Bentley et al., 1993). However, owing to the large interindividual variation in peroxisome morphometrics observed in these studies, together with cell to cell variations and lobular variations, it is difficult to attach any clear biological significance to these findings (Bentley et al., 1993). Generally, peroxisome proliferators have not been reported to produce any significant effects on marker enzyme activities and/or peroxisomes in cultured primate and human hepatocytes (Bieri, 1993; Bentley et al., 1993; Foxworthy and Eacho, 1994). Some studies have also examined species differences in effects on cell replication. Both nafenopin and Wy-14,643 have been reported to stimulate replicative DNA synthesis in rat, but not in Syrian hamster, hepatocytes (Price et al., 1992; Lake et al., 1993). Although peroxisome proliferators can stimulate DNA synthesis in cultured rat hepatocytes, methylclofenapate was reported to be ineffective in guinea pig, marmoset and human hepatocytes (Elcombe and Styles, 1989). Similarly, nafenopin has also been reported not to induce replicative DNA synthesis in human hepatocytes (Parzefall et al., 1991). Risk assessment of rodent liver peroxisome proliferators The key issues concerning the risk assessment of rodent liver peroxisome proliferators include: (a) Genotoxicity. (b) Likely human exposure. (c) Compound potency and no effect levels. (d) Precise mechanism(s) of liver tumour formation. (e) Species differences in response. Generally, peroxisome proliferators are considered to be non-genotoxic agents (Bentley et al., 1993; Budroe and Williams, 1993) and hence should be assessed differently from genotoxic carcinogens (Weisburger, 1994). Human exposure to rodent peroxisome proliferators depends on the

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intended usage of the particular compound. While hypolipidaemic agents are only administered to a restricted population of humans, exposure to industrial chemicals such as plasticisers is obviously far more widespread. For example, based on food surveillance surveys the daily human exposure to DEHA was reported to be 16 and 8.2 mg per person per day in 1987 and 1990, respectively (MAFF 1987, 1990). In another study, where DEHA intake was assessed by measuring urinary levels of the major metabolite 2-ethylhexanoic acid, a median value of 2.7 mg per person per day was reported (Loftus et al., 1994). Apart from likely human exposure, consideration should be made of the relative potency of the particular compound to produce peroxisome proliferation and liver tumours in rodents. Plasticisers such as DEHP and DEHA are far less potent than certain therapeutic agents and experimentally used compounds (Reddy et al., 1986; Barber et al., 1987; Bentley et al., 1993; Lake and Lewis, 1993). Moreover rodent liver peroxisome proliferators exhibit clear no effect levels for both peroxisome proliferation and for tumour formation. For example, in the rat no effect levels for liver tumour formation have been observed in studies with several compounds including bezafibrate, clofibrate, DEHA and DEHP (Hartig et al., 1982; NTP 1982a, b). In addition, the threshold for tumour formation in rodents is appreciably greater than the threshold for peroxisome proliferation (Hartig et al., 1982; Reddy et al., 1986; Bentley et al., 1993). Several mechanisms have been proposed to account for why peroxisome proliferators produce tumours in rodent liver. If these various hypotheses are combined then a role for increased cell replication in peroxisome proliferatorinduced hepatocarcinogenesis may be readily identified. For example, if hepatocytes are transformed by either oxidative stress-induced damage or by alternative mechanisms, such initiated cells may be promoted to liver tumours by enhanced cell replication. Certainly peroxisome proliferators are effective promoters of certain populations of initiated cells and recent studies suggest that peroxisome proliferators can influence rates of both cell replication and cell death in particular populations of hepatocytes (Grasl-Kraupp et al., 1993; Popp and Cattley, 1993; Marsman and Popp, 1994). With respect to species differences, rats and mice are clearly responsive species, whereas the majority of both in vivo and in vitro studies suggest that primates including man are either essentially refractory or certainly much less responsive to rodent peroxisome proliferators. However while effects on peroxisome morphology and marker enzyme activities have been extensively studied, few investigations have examined species differences in peroxisome proliferator-induced cell replication and liver tumour formation. As enhanced cell replication appears to play a role in peroxisome proliferator-induced hepatocarcinogenesis in rats and mice, it would

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appear to be an important biomarker for assessing species differences in response. Rodent peroxisome proliferators do not appear to stimulate replicative DNA synthesis in vivo in Syrian hamster hepatocytes and in vitro in human hepatocytes (Elcombe and Styles 1989; Parzefall et al., 1991; Price et al., 1992; Lake et al., 1993). With respect to tumour formation, nafenopin and Wy-14,643 (two potent peroxisome proliferators) were reported not to produce liver lesions in the Syrian hamster although both compounds produced liver nodules and hepatocellular carcinoma after 60 weeks in the rat (Lake et al., 1993). Similarly, clofibrate was reported not to increase liver weight or produce liver tumours in marmosets after 6.5 years treatment (Tucker and Orton, 1993) and in an ongoing study ciprofibrate was found not to produce any morphological changes in marmoset liver after 3 years administration (Graham et al., 1994). In conclusion, the present literature suggests that rodent peroxisome proliferators are non-genotoxic agents which should be assessed differently from genotoxic compounds for human hazard (Weisburger, 1994). Assessment of likely human exposure and compound potency are also important factors together with information on compound no effect levels and evidence of species differences in response. Rodent liver peroxisome proliferators as a class of chemicals thus do not appear to pose any serious hazard for man. However, it would be desirable to elucidate further the mechanism(s) of peroxisome proliferator-induced hepatocarcinogenesis in susceptible species (i.e. the rat and mouse). From such studies the most appropriate biomarkers of liver tumour formation could be identified and examined in studies of species differences possibly including in vitro studies with human hepatocytes. Finally, further carcinogenicity studies in partially responsive (e.g. Syrian hamster) and non-responsive (e.g. guinea pig) species would strengthen the conclusion that peroxisome proliferators do not constitute any significant hazard to man. Acknowledgement We thank the UK Ministry of Agriculture, Fisheries and Food for financial support of BIBRA studies on hepatic peroxisome proliferation. References ALBRO, P.W., CORBETT, J.T., SCHROEDER, J.L. and JORDAN, S.T., 1983, Incorporation of radioactivity from labelled di-(2-ethylhexyl)phthalate into DNA of rat liver in vivo. Chemico-Biological Interactions, 44, 1–16. BARBER, E.D., ASTILL, B.D., MORAN, E.J., SCHNEIDER, B.F., GRAY, T.J. B., LAKE, B.G. and EVANS, J.G., 1987, Peroxisome induction studies on seven phthalate esters, Toxicology and Industrial Health, 3, 7–22.

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BENTLEY, P., CALDER, I., ELCOMBE, C., GRASSO, P., STRINGER, D. and WIEGAND, H.-J., 1993, Hepatic peroxisome proliferation in rodents and its significance for humans, Food and Chemical Toxicology, 31, 857–907. BIEBER, L.L., KRAHLING, J.B., CLARKE, P.R.H., VALKNER, K.J. and TOLBERT, N.E., 1981, Carnitine acyltransferases in rat liver peroxisomes, Archives of Biochemistry and Biophysics, 211, 599–604. BIERI, F., 1993, Cultured hepatocytes: a useful in vitro system to investigate effects induced by peroxisome proliferators and their species specificity, in Gibson, G. and Lake, B. (Eds) Peroxisomes: Biology and Importance in Toxicology and Medicine, pp. 299–311, London: Taylor and Francis. BUDROE, J.D. and WILLIAMS, G.M. 1993, Genotoxicity studies of peroxisome proliferators, in Gibson, G. and Lake, B. (Eds) Peroxisomes: Biology and Importance in Toxicology and Medicine, pp. 525–68, London: Taylor and Francis. CATTLEY, R.C. and POPP, J.A., 1989, Differences between the promoting activities of the peroxisome proliferator WY-14,643 and phenobarbital in rat liver, Cancer Research, 49, 3246–51. CATTLEY, R.C., MARSMAN, D.S. and POPP, J.A., 1991, Age-related susceptibility to the carcinogenic effect of the peroxisome proliferator WY-14, 643 in rat liver, Carcinogenesis, 12, 469–73. COHEN, S.M. and ELLWEIN, L.B., 1990, Cell proliferation in carcinogenesis, Science, 249, 1007–11. COHEN, S.M. and ELLWEIN, L.B., 1991, Genetic errors, cell proliferation and carcinogenesis, Cancer Research, 51, 6493–505. COHEN, A.J. and GRASSO, P., 1981, Review of the hepatic response to hypolipidaemic drugs in rodents and assessment of its toxicological significance to man, Food and Chemical Toxicology, 19, 585–605. DIRVEN, H.A.A.M., VAN DEN BROEK, P.H.H., PETERS, J.G.P., NOORDHOEK, J. and JONGENEELEN, F.J., 1992, Microsomal lauric acid hydroxylase activities after treatment of rats with three classical cytochrome P-450 inducers and peroxisome proliferating compounds, Biochemical Pharmacology, 43, 2621–9. EACHO, P.I., FOXWORTHY, P.S., JOHNSON, W.D., HOOVER, D.M. and WHITE, S.L., 1986, Hepatic peroxisomal changes induced by a tetrazolesubstituted alkoxyacetophenone in rats and comparison with other species, Toxicology and Applied Pharmacology, 83, 430–7. EACHO, P.I., LANIER, T.L. and BRODHECKER, C.A., 1991, Hepatocellular DNA synthesis in rats given peroxisome proliferating agents: comparison of Wy-14,643 to clofibric acid, nafenopin and LY 171883, Carcinogenesis, 12, 1557– 61. ELCOMBE, C.R., 1985, Species differences in carcinogenicity and peroxisome proliferation due to trichloroethylene: a biochemical human hazard assessment, Archives of Toxicology, Supplement 8, 6–17. ELCOMBE, C.R. and MITCHELL, A.M., 1986, Peroxisome proliferation due to di (2-ethylhexyl)phthalate (DEHP): species differences and possible mechanisms, Environmental Health Perspectives, 70, 211–19. ELCOMBE, C.R. and STYLES, J.A., 1989, Species differences in peroxisome proliferation, Toxicologist, 9, 63.

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FOXWORTHY, P.S. and EACHO, P.I., 1994, Cultured hepatocytes for studies of peroxisome proliferation: methods and applications, Journal of Pharmacological and Toxicological Methods, 31, 21–30. GIBSON, G.G., 1989, Comparative aspects of the mammalian cytochrome P450IV gene family, Xenobiotica, 19, 1123–48. GOEL, S.K., LALWANI, N.D., FAHL, W.E. and REDDY, J.K., 1985, Lack of covalent binding of peroxisome proliferators nafenopin and WY-14,643 to DNA in vivo and in vitro, Toxicology Letters, 24, 37–43. GRAHAM, M.J., WILSON, S.A., WINHAM, M.A., SPENCER, A.J., REES, J.A., OLD, S.L. and BONNER, F.W., 1994, Lack of peroxisome proliferation in marmoset liver following treatment with ciprofibrate for 3 years, Fundamental and Applied Toxicology, 22, 58–64. GRASL-KRAUPP, B., HUBER, W. and SCHULTE-HERMANN, R., 1993, Are peroxisome proliferators tumour promoters in rat liver? in Gibson, G. and Lake, B. (Eds) Peroxisomes: Biology and Importance in Toxicology and Medicine, pp. 667–93, London: Taylor and Francis. GRAY, T.J.B., BEAMAND, J.A., LAKE, B.G., FOSTER, J.R. and GANGOLLI, S.D., 1982, Peroxisome proliferation in cultured rat hepatocytes produced by clofibrate and phthalate ester metabolites, Toxicology Letters, 10, 273–9. GUPTA, R.C., GOEL, S.K., EARLEY, K., SINGH, B. and REDDY, J.K., 1985, 32PPostlabelling analysis of peroxisome proliferator-DNA adduct formation in rat liver in vivo and hepatocytes in vitro, Carcinogenesis, 6, 933–6. HANEFELD, M., KEMMER, C. and KADNER, E., 1983, Relationship between morphological changes and lipid-lowering action of pchlorophenoxyisobutyric acid (CPIB) on hepatic mitochondria and peroxisomes in man, Atherosclerosis, 46, 239–46. HARTIG, F., STEGMEIER, K., HEBOLD, G.,ÖZEL, M. and FAHIMI, H.D., 1982, Study of liver enzymes: peroxisome proliferation and tumor rates in rats at the end of carcinogenicity studies with bezafibrate and clofibrate, Annals of the New York Academy of Science, 386, 464–7. HAWKINS, J.M., JONES, W.E., BONNER, F.W. and GIBSON, G.G., 1987, The effect of peroxisome proliferators on microsomal, peroxisomal and mitochondrial enzyme activities in the liver and kidney, Drug Metabolism Reviews, 18, 441–515. HESS, R., STÄUBLI, W. and REISS, W., 1965, Nature of the hepatomegalic effect produced by ethyl chlorophenoxyisobutyrate in the rat, Nature, 208, 856–8. HINTON, R.H. and PRICE, S.C., 1993, Extrahepatic peroxisome proliferation and the extrahepatic effects of peroxisome proliferators, in Gibson, G. and Lake, B. (Eds) Peroxisomes: Biology and Importance in Toxicology and Medicine, pp. 487– 511, London: Taylor and Francis. HWANG, J.-J., HSAI, M.T.S. and JIRTLE, R.L., 1993, Induction of sister chromatid exchange and micronuclei in primary cultures of rat and human hepatocytes by the peroxisome proliferator Wy-14,643, Mutation Research, 286, 123–33. ISHII, H., FUKUMORI, N., HORIE, S. and SUGA, T., 1980, Eifects of fat content in the diet on hepatic peroxisomes of the rat, Biochimica et Biophysica Acta, 617, 1–11.

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KRAUPP-GRASL, B., HUBER, W., TAPER, H. and SCHULTE-HERMANN, R., 1991, Increased susceptibility of aged rats to hepatocarcinogenesis by the peroxisome proliferator nafenopin and the possible involvement of altered liver foci occurring spontaneously, Cancer Research, 51, 666–71. LAKE, B.G., 1993, Role of oxidative stress and enhanced cell replication in the hepatocarcinogenicity of peroxisome proliferators, in Gibson, G. and Lake, B. (Eds) Peroxisomes: Biology and Importance in Toxicology and Medicine, pp. 595– 618, London: Taylor and Francis. LAKE, B.G. and LEWIS, D.F.V., 1993, Structure-activity relationships for chemically induced peroxisome proliferation in rodent liver, in Gibson, G. and Lake, B. (Eds) Peroxisomes: Biology and Importance in Toxicology and Medicine, pp. 313– 42, London: Taylor and Francis. LAKE, B.G., GRAY, T.J.B., PELS RIJCKEN, W.R., BEAMAND, J.A. and GANGOLLI, S.D., 1984a, The effect of hypolipidaemic agents on peroxisomal β-oxidation and mixed-function oxidase activities in primary cultures of rat hepatocytes. Relationship between induction of palmitoyl-CoA oxidation and lauric acid hydroxylation, Xenobiotica, 14, 269–76. LAKE, B.G., GRAY, T.J.B., FOSTER, J.R., STUBBERFIELD, C.R. and GANGOLLI, S.D., 1984b, Comparative studies on di-(2-ethylhexyl)phthalateinduced hepatic peroxisome proliferation in the rat and hamster, Toxicology and Applied Pharmacology, 72, 46–60. LAKE, B.G., GRAY, T.J.B. and GANGOLLI, S.D., 1986, Hepatic effects of phthalate esters and related compounds—in vivo and in vitro correlations, Environmental Health Perspectives, 67, 283–90. LAKE, B.G., LEWIS, D.F.V. and GRAY, T.J.B., 1988, Structure-activity relationships for peroxisome proliferation, Archives of Toxicology (Supplement 12), 217–24. LAKE, B.G., EVANS, J.G., GRAY, T.J.B., KÖRÖSI, S.A. and NORTH, C.J., 1989, Comparative studies on nafenopin-induced hepatic peroxisome proliferation in the rat, Syrian hamster, guinea pig and marmoset, Toxicology and Applied Pharmacology, 99, 148–60. LAKE, B.G., EVANS, J.G., CUNNINGHAME, M.E. and PRICE, R.J., 1993, Comparison of the hepatic effects of nafenopin and Wy-14,643 on peroxisome proliferation and cell replication in the rat and Syrian hamster, Environmental Health Perspectives 101 (Supplement 5), 241–8. LALWANI, N.D., REDDY, M.K., GHOSH, S., BARNARD, S.D., MOLELLO, J. A. and REDDY, J.K., 1985, Induction of fatty acid β-oxidation and peroxisome proliferation in the liver of Rhesus monkeys by DL-040, a new hypolipidemic agent, Biochemical Pharmacology, 34, 3473–82. LAZAROW, P.B. and DEDUVE, C., 1976, A fatty acyl-CoA oxidizing system in rat liver peroxisomes: enhancement by clofibrate, a hypolipidaemic drug, Proceedings of the National Academy of Sciences USA, 73, 2043–6. LIN, L.I.-K., 1987, The use of multivariate analysis to compare peroxisome induction data on phthalate esters in rats, Toxicology and Industrial Health, 3, 25–47. LOCK, E.A., MITCHELL, A.M. and ELCOMBE, C.R., 1989, Biochemical mechanisms of induction of hepatic peroxisome proliferation, Annual Reviews of Pharma-cology and Toxicology, 29, 145–63.

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LOFTUS, N.J., WOOLLEN, B.H., STEEL, G.T., WILKS, M.F. and CASTLE, L., 1994, An assessment of the dietary uptake of di-2-(ethylhexyl)adipate (DEHA) in a limited population study, Food and Chemical Toxicology, 32, 1–5. MAKOWSKA, J.M., GIBSON, G.G. and BONNER, F.W., 1992, Species differences in ciprofibrate induction of hepatic cytochrome P450 4A1 and peroxisome proliferation, Journal of Biochemical Toxicology, 7, 183–91. MARSMAN, D.S. and POPP, J.A., 1994, Biological potential of basophilic hepatocellular foci and hepatic adenoma induced by the peroxisome proliferator, Wy-14,643, Carcinogenesis, 15, 111–17. MARSMAN, D.S., CATTLEY, R.C., CONWAY, J.G. and POPP, J.A., 1988, Relationship of hepatic peroxisome proliferation and replicative DNA synthesis to the hepatocarcinogenicity of the peroxisome proliferators di(2ethylhexyl)phthalate and [4-chloro-6-(2,3-xylidino)-2-pyrimidinylthio]acetic acid (Wy-14,643) in rats, Cancer Research, 48, 6739–44. Ministry of Agriculture, Fisheries and Food (MAFF), 1987, Survey of Plasticiser Levels in Food Contact Materials and in Foods, Food Surveillance Paper No. 21, London: HMSO. Ministry of Agriculture, Fisheries and Food (MAFF), 1990, Survey of Plasticiser Levels in Food Contact Materials and in Foods, Food Surveillance Paper No. 30, London: HMSO. MOODY, D.E., REDDY, J.K., LAKE, B.G., POPP, J.A. and REESE, D.H., 1991, Peroxisome proliferation and nongenotoxic carcinogenesis: commentary on a symposium, Fundamental and Applied Toxicology, 16, 233–48. NCI, 1976, Carcinogenesis of trichloroethylene (CAS No. 79–01–6), NCI-CGTR-2, National Cancer Institute. NTP (National Toxicology Program), 1982a, Carcinogenesis bioassay of di(2-ethylhexyl)phthalate (CAS No. 117–81–7) in F344 rats and B6C3F1 mice (feed study), Technical Report Series No. 217. NTP (National Toxicology Program), 1982b, Carcinogenesis bioassay of di(2ethyl- hexyl)adipate (CAS No. 103–23–1) in F344 rats and B6C3F1 mice (feed study), Technical Report Series No. 212. ORTON, T.C., ADAM, H.K., BENTLEY, M., HOLLOWAY, B. and TUCKER, M. J., 1984, Clobuzarit: species differences in the morphological and biochemical response of the liver following chronic administration, Toxicology and Applied Pharmacology, 73, 138–51. OSUMI, T. and HASHIMOTO, T., 1978, Enhancement of fatty acyl-CoA oxidizing activity in rat liver peroxisomes by di(2-ethylhexyl)phthalate, Journal of Biochemistry, 83, 1361–5. PAGET, G.E., 1963, Experimental studies of the toxicity of Atromid with particular reference to fine structural changes in the livers of rodents, Journal of Atherosclerosis Research, 3, 729–37. PARZEFALL, W., ERBER, E., SEDIVY, R. and SCHULTE-HERMANN, R., 1991, Testing for induction of DNA synthesis in human hepatocyte primary cultures by rat liver tumor promoters, Cancer Research, 51, 1143–7. POPP, J.A. and MARSMAN, D.S., 1991, Chemically-induced cell proliferation in liver carcinogenesis, in Butterworth, B.E., Slaga, T.J., Farland, W. and McClain, M. (Eds) Chemically Induced Cell Proliferation: Implications for Risk Assessment, pp. 389–95, New York: Wiley-Liss.

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POPP, J.A. and CATTLEY, R.C., 1993, Peroxisome proliferators as initiators and promoters of rodent hepatocarcinogenesis, in Gibson, G. and Lake, B. (Eds) Peroxisomes: Biology and Importance in Toxicology and Medicine, pp. 653– 65, London: Taylor and Francis. PRICE, R.J., EVANS, J.G. and LAKE B.G., 1992, Comparison of the effects of nafenopin on hepatic peroxisome proliferation and replicative DNA synthesis in the rat and Syrian hamster, Food and Chemical Toxicology 30, 937–44. REDDY, J.K. and LALWANI, N.D., 1983, Carcinogenesis by hepatic peroxisome proliferators: evaluation of the risk of hypolipidemic drugs and industrial plasticisers to humans, CRC Critical Reviews in Toxicology, 12, 1–58. REDDY, J.K. and RAO, M.S., 1989, Oxidative DNA damage caused by persistent peroxisome proliferation: its role in hepatocarcinogenesis, Mutation Research, 214, 63–8. REDDY, J.K., LALWANI, N.D., QURESHI, S.A., REDDY, M.K. and MOEHLE, C.M., 1984, Induction of hepatic peroxisome proliferation in non-rodent species, including primates, American Journal of Pathology, 114, 171–83. REDDY, J.K., REDDY, M.K., USMAN, M.I., LALWANI, N.D. and RAO, M. S., 1986, Comparison of hepatic peroxisome proliferative effect and its implication for hepatocarcinogenicity of phthalate esters, di(2-ethylhexyl) phthalate and di(2-ethylhexyl)adipate with a hypolipidemic drug, Environmental Health Perspectives, 65, 317–27. RHODES, C.ORTON, T.C., PRATT, I.S., BATTEN, P.L., BRATT, H., JACKSON, S.J. and ELCOMBE, C.R., 1986, Comparative pharmacokinetics and subacute toxicity of di(2-ethylhexyl)phthalate (DEHP) in rats and marmosets: extrapolation of effects in rodents to man, Environmental Health Perspectives, 65, 299–308. RODRICKS, J.V. and TURNBULL, D., 1987, Interspecies differences in peroxisomes and peroxisome proliferation, Toxicology and Industrial Health, 3, 197–212. SCHULTE-HERMANN, R., TIMMERMANN-TROSIENER, I. and SCHLUPPLER, J., 1983, Promotion of spontaneous preneoplastic cells in rat liver as a possible expla nation of tumor production by nonmutagenic compounds, Cancer Research, 43, 839–44. SCHULTE-HERMANN, R., KRAUPP-GRASL, B., BURSCH, W., GERBRACHT, U. and TIMMERMANN-TROSIENER, I., 1989, Effects of non-genotoxic hepatocarcinogens phenobarbital and nafenopin on phenotype and growth of different populations of altered foci in rat liver, Toxicologic Pathology, 17, 642–50. SHARMA, R., LAKE, B.G., FOSTER, J. and GIBSON, G.G., 1988a, Microsomal cytochrome P-452 induction and peroxisome proliferation by hypolipidaemic agents in rat liver. A mechanistic inter-relationship, Biochemical Pharmacology, 37, 1193–201. SHARMA, R., LAKE, B.G. and GIBSON, G.G., 1988b, Co-induction of microsomal cytochrome P-452 and the peroxisomal fatty acid β-oxidation pathway in the rat by clofibrate and di-(2-ethylhexyl)phthalate. Dose-response studies, Biochemical Pharmacology, 37, 1203–6.

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STOTT, W.T., 1988, Chemically induced proliferation of peroxisomes: implications for risk assessment, Regulatory Toxicology and Pharmacology, 8, 125–59. TAKAGI, A., SAI, K., UMEMURA, T., HASEGAWA, R. and KUROKAWA, Y., 1990, Significant increase of 8-hydroxydeoxyguanosine in liver DNA of rats following short-term exposure to the peroxisome proliferators di(2-ethylhexyl) phthalate and di(2-ethylhexyl)adipate, Japanese Journal of Cancer Research, 81, 213–15. THORNBER, C.W., 1979, Isosterism and molecular modification in drug design, Chemical Society Reviews, 8, 563–80. TUCKER, M.J. and ORTON, T.C., 1993, Toxicological studies in primates with three fibrates, in Gibson, G. and Lake, B. (Eds) Peroxisomes: Biology and Importance in Toxicology and Medicine, pp. 425–47, London: Taylor and Francis. VON DÄNIKEN, A., LUTZ, W.K. and SCHLATTER, C., 1981, Lack of covalent binding to rat liver DNA of the hypolipidemic drugs clofibrate and fenofibrate, Toxicology Letters, 7, 311–19. VON DÄNIKEN, A., LUTZ, W.K., JÄCKH, R. and SCHLATTER, C., 1984, Investi-gation of the potential for binding of di(2-ethylhexyl)phthalate (DEHP) and di(2-ethylhexyl)adipate (DEHA) to liver DNA in vivo, Toxicology and Applied Pharmacology, 73, 373–87. WADA, N., MARSMAN, D.S. and POPP, J.A., 1992, Dose-related effects of hepatocarcinogen, Wy-14,643 on peroxisomes and cell replication, Fundamental and Applied Toxicology, 18, 149–54. WEISBURGER, J.H., 1994, Does the Delaney Clause of the U.S. Food and Drug Laws prevent human cancers? Fundamental and Applied Toxicology, 22, 483– 93.

18 Neurotoxicity Testing of Industrial Compounds: in vivo Markers and Mechanisms of Action KORNELIS J.VAN DEN BERG,1 JAN-BERT P.GRAMSBERGEN,2 ELISABETH M.G.HOOGENDIJK,1 JAN H.C.M.LAMMERS,1 WILLEM S.SLOOT1 and BEVERLY M.KULIG1 1

TNO Nutrition and Food Research Institute, Rijswijk; 2 Erasmus University, Rotterdam Introduction

Neurotoxicity assessment is designed to provide an answer to the question of whether or not a particular chemical is able to evoke some form of adverse effect specifically associated with the nervous system. The development of risk assessment procedures is a long-term goal in view of the complexity of the target organ involved, e.g. the nervous system. As risk identification is a first step in the risk assessment paradigm of chemicals (NAS, 1983), for neuro-toxicity this translates at present into procedures and methods aimed at characterization of altered neurobehaviour of exposed experimental animals and abnormalities in the morphology of the nervous system. It has been suggested that risk assessment procedures may take more the approach of ‘exposuredoseresponse’ (Andersen, 1991) in which mechanisms play an important role at various levels, e.g. from disposition of the chemical through the body to target tissues, via biochemical interactions at the molecular level to a toxic response as altered behaviour. Understanding the mechanisms involved in this sequence of events is helpful to arrive in the future at quantitative risk assessment procedures, for instance biologically-based modelling (Borghoff et al., 1991). In addition, a better knowledge of the mechanistic principles of neuro-toxic agents may lead to the development of specific biomarkers that would further improve the efficiency of procedures for neurotoxicity screening, given the magnitude of the number of potentially neurotoxic compounds (NRC, 1992). Only for a small select group of industrial chemicals are mechanisms of neurotoxic action more or less characterized. Perhaps the oldest examples are the class of organophosphate (OP) pesticides where the molecular targets have been identified as acetylcholinesterase (AChE) and neuropathy target esterase (NTE), the latter being associated with organophosphate-

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induced delayed neuropathy (OPIDN) (Cherniack, 1988). This has led to investigations into structure-activity relationships of OPs with NTE in vitro that have proved to be valuable in predicting OPIDN (Davis et al. 1985). Furthermore, certain tissue culture systems, such as neuroblastoma cell lines, contain NTE and AChE activity that may be suited for identification of OPs causing OPIDN (Veronesi, 1992). Biochemical assays for NTE and AChE, in conjunction with neurobehavioural observations and neuropathology, are currently incorporated into US and Japanese neurotoxicity testing guidelines. A second group of industrial chemicals for which indications exist on their mechanism of action include certain compounds causing peripheral neuro-pathy such as acrylamide, carbon disulphide and n-hexane. It is generally assumed that the primary action of these chemicals is the crosslinking of axonal proteins (Graham et al., 1982), a process that blocks axonal transport (Sickles, 1991) and may lead to degeneration of distal axons (Spencer and Schaumberg, 1980). The basic information thus collected is yet to be developed into a useful biomarker. The pyrethroid insecticides represent a third group of chemicals for which the neurotoxic mechanism of action has been elucidated. The major symptoms of pyrethroid intoxication, e.g. convulsions, tremors, paralysis, are primarily the result of interaction of the pyrethroids with sodiumchannels on nerve membranes (Lund and Narahashi, 1982). Whereas under normal circumstances a sodium channel is only opened during a depolarization event, pyrethroids prolong opening of these channels thereby causing repetitive excitation of nerve and nerve terminals. In tissue culture systems, e.g. neuroblastoma cell lines, direct electrophysiological studies on cells have been done to investigate the effectiveness of different pyrethroids for opening of sodium channels (Oortgiesen et al., 1989). Biomarkers of neurotoxicity It is clear that a mechanistic understanding of the neurotoxicity of suspected chemicals is and will be proceeding at a slow pace. Eventual utilization of this knowledge in the form of biological markers for neurotoxicity risk assessment procedures is, necessarily, a long-term objective. In the mean time alternative approaches have been proposed recently that (a) may provide biomarkers that could aid to further define underlying mechanisms and (b) are directly linked to neurotoxic mechanisms of actions. Gliotypic and neurotypic proteins Insults to the brain by a large variety of agents or conditions, e.g. viral infection, auto-immune encephalitis, trauma and chemicals, evoke a fairly

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stereo typic response of astrocytes (Eng, 1988). The normally rather quiescent astrocytes change into reaction astrocytes with a characteristic morphological appearance, a process known as reactive gliosis or astrogliosis. It has been proposed that this fairly universal astrocytic reaction might be a useful parameter for risk identification purposes (O’Callaghan, 1992; Rosengren and Haglid, 1989). The hypertrophic response of astrocytes is biochemically characterized by a greatly enhanced synthesis of glial fibrillary acidic protein (GFAP), a major structural protein of intermediate filaments (Eng, 1988). Under a variety of experimentallyinduced insults to the central nervous system, increased GFAP immunoreactivity has consistently been found in association with neuronal damage (Eng, 1988; O’Callaghan and Miller, 1989). It must be kept in mind, however, that reactive astrogliosis represents an indirect indication of nerve cell damage or loss. A strategy has been proposed (Brock and O’Callaghan, 1987) to collect quantitative data on astrogliosis, for example for GFAP concentrations, in association with information on changes of neuron-specific proteins. e.g. synapsin I or synaptophysin. The general idea is that enhanced GFAP levels in combination with decreased synaptophysin concentrations in the brain would be strongly indicative of a neurotoxic event. Recent methodological developments have led to quantitative procedures for assessment of GFAP and synaptophysin in nerve tissues by dot-blot immunoassay and ELISA (Jahn et al. 1984; O’Callaghan, 1991). Cerebral calcium accumulation Nerve cells, like most animal cells, possess a complicated system comprising calcium gates, pumps and channels to maintain free cytosolic calcium ion levels within a low physiological range (Pounds, 1990). Under pathological conditions, including hypoxia-ischaemia and status epilepticus, calcium overload in vulnerable neurons has been observed that is thought to be associated with the process of cell death (both necrosis and apoptosis) (Boobis et al. 1989; Siesjo and Bengtsson, 1989). Cerebral calcium accumulation has, therefore, been proposed as a potential index of brain pathology (Korf et al., 1986). Free radical formation Free radicals may be involved in mechanisms of toxicity, including neurotoxicity of chemicals and are gaining more attention as is obvious from a number of recent reviews on this topic (LeBel and Bondy, 1991; Halliwell et al. 1992; Aust et al., 1993). Basically, free radicals are molecules that contain one or more unpaired electrons, whereas most molecules are nonradicals. Because of the unpaired electron(s) free radicals are highly

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reactive chemical intermediates. Once a free radical is formed it can initiate a chain of reactions as the free electron is passed from one molecule to the other. In vivo, free radicals involving oxygen species are continuously produced and evolution has also provided the body with defence mechanisms such as superoxide dismutase (SOD), catalase, and scavengers such as glutathione, ascorbate, etc. Oxidative stress by free radicals may be an important neuropathological mediator following exposure to a number of neurotoxic agents (LeBel and Bondy, 1991). Examples of neurotoxic chemicals suspected of causing free radicals include chlordecone, ethanol, methamphetamine, methyl mercury, toluene, triethyl lead and trimethyltin. Of special interest is a ‘designer’ drug called MPTP1 causing destruction of dopaminergic neurons of the basal ganglia with symptoms similar to Parkinson’s disease. It is the neurotoxic metabolite MPP+2 that has been found to induce cerebral oxygen radical formation in vitro. There is also suggestive evidence to indicate that free radical scavengers may provide protection of the basal ganglia against neuro-toxic effects of MPP+ (LeBel and Bondy, 1991). The occurrence of these kinds of compounds has led, among other things, one to suspect possible involvement of environmental chemicals and factors associated with diet and/or lifestyle in Parkinson’s disease (Russell, 1992; Semchuk et al., 1993). Model neurotoxins In order to validate an approach based on changes in these proposed biomarkers experimental studies were performed using various model neuro-toxicants, including trimethyltin, kainic acid, heavy metals such as lead, methylmercury and manganese, and developmental neurotoxicants, e.g. polychlorinated biphenyls. Trimethyltin Trimethyltin (TMT) is known to cause in adult rats a rather selective neuronal degeneration in specific regions of the brain, notably in limbic structures such as the hippocampus where extensive loss of pyramidal cells in CA fields are observed by standard histopathological procedures (O’Callaghan, 1988). A single systemic dose of TMT (7.5 mg kg−1) given to adult rats caused, after a period of 3 weeks, an approximately three-fold enhanced level of GFAP in hippocampus (Figure 18.1, upper left panel). In a number of other brain regions, e.g. different parts of the cortex, thalamus, striatum, cerebellum and brain stem, no significant changes in GFAP were observed. Assessment of synaptophysin, a structural protein of synaptic vesicles of neurons (Jahn et al. 1985) in the same brain regions is given in Figure 18.1 (lower left panel). TMT induced a significant

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Figure 18.1 Changes in cerebral glial fibrillary acidic protein (GFAP) and synaptophysin concentrations by trimethyltin (TMT) and kainic acid. Results are expressed as mean±SEM. Open bars, control; closed bars, 7.5 mg trimethyltin (TMT) per kg or 12 mg kainic acid per kg; HP, hippopcampus; AM, amygdala; *, p<0.05.

reduction of synaptophysin levels by approximately 30 per cent in the hippocampus.

1 2

l-Methyl-4-phenyl-1,2,3,6-tetrahydropyridine. 1-Methyl-4-phenyl-pyridinium.

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Kainic acid Kainic acid (KA), a glutamate agonist, is a potent model neurotoxin that causes neurodegeneration in a number of limbic structures including hippocampus, amygdala, piriform cortex (Sperk et al. 1983; Ben-Ari, 1985). It is generally assumed that the mechanism of action involves interaction of KA with a special class of glutamate receptors and release of endogenous excitatory amino acids in levels that are detrimental to neurons (Meldrum and Garthwaite, 1990). Kainic acid did induce, in a single systemic dose (12 mg kg−1), highly increased concentrations of GFAP in a number of target structures. For instance in hippocampus and amygdala GFAP levels did rise to 650 and 960 per cent of control levels (Figure 18.1, upper right panel). Recent results from this laboratory have revealed that in a time-course study (Van den Berg and Gramsbergen, 1993) maximum levels are obtained after 4 weeks that remained highly elevated in most target brain regions for at least a period of 6 months (Figure 18.2). The quantitative data on GFAP concentrations were supported by increased GFAP immunoreactivity in hippocampal sections visualized by immunohistochemical procedures. In addition to the damage in the hippocampus, permanently enhanced GFAP levels were found in other brain regions, e.g. piriform cortex, septum (Gramsbergen and Van den Berg, 1994) known to be targets of KA neurotoxicity. Synaptophysin levels were significantly reduced by KA in hippocampus and amygdala to a comparable degree (Figure 18.1, lower right panel). These data indicate a decrease in synaptophysin content encountered in brain regions where neuronal elements are known to be lost by these model neurotoxins. The magnitude of the changes in synaptophysin concentrations were much smaller than those of GFAP in the same brain structures. This may be explained by the fairly selective neuronal loss in specific layers of, for instance, the hippocampus, by TMT and KA. The effect is thus rather diluted in a biochemical procedure. Once an effect is scored, more detailed biochemical analysis is possible by using a punch technique after microdissection of brain nuclei (Palkovits and Brownstein, 1988). Alternatively, a follow-up by histopathological procedures, for example by the cupric silver degeneration stain, would provide further details of neuronal damage (O’Callaghan and Jensen, 1992). In the experiments described above there was no clear correlation between the graded regional GFAP response and decrease of synaptophysin concentration. This may suggest a differential region-specific response of astrocytes towards neuronal injury. In our laboratory cerebral calcium accumulation was determined recently after a single systemic dose of KA (12 mg kg−1, i.p.), given to adult rats. A rapid uptake of 45Ca was observed in various regions of the brain. A

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Figure 18.2 Time-course of GFAP concentration in hippocampus by kainic acid. Rats were treated with a single systemic dose of kainic acid (12 mg kg−1, i.p.). GFAP was determined by ELISA at the times indicated. Results are expressed as mean±SEM (from Van den Berg and Gramsbergen, 1993, with permission).

time-course study, covering a period of 6 months, has indicated that in the hippocampus a peak of 45Ca uptake occurred already after 4 days and normal values were reached only after another 2–4 months (Van den Berg and Gramsbergen, 1993). More recent results indicate that other limbic areas display similar kinetics, while 45Ca uptake in striatum and cortex return faster to normal values, e.g. within 2 weeks. Only in the thalamus a long-term sustained 45Ca uptake was present for a period of 6 months. In general a good correlation was found between the regions showing sustained 45Ca accumulation and those known to be targets of KA neurotoxicity. When these data on 45Ca accumulation were related to effects of KA on GFAP concentrations, also a good agreement was observed concerning the brain regions involved (Gramsbergen and Van den Berg, 1994). Histopathological examination of hippocampal sections revealed extensive neurodegeneration by KA in CA1, CA3 and CA4 regions (Van den Berg and Gramsbergen, 1993). Heavy metals Lead may produce in children symptoms of acute encephalopathy after exposure to high doses and at lower dose levels learning disorders and

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hyperactive behaviour. In adults neurotoxicity is more often encountered as peripheral neuropathy after chronic occupational exposure to lead (Marsh, 1985). A number of mechanisms have been implicated in lead neurotoxicity (Bressler and Goldstein, 1991), but as yet no central hypothesis has emerged. In order to investigate the effect of lead on CNS structural proteins, a subchronic dosing experiment with adult rats was performed in which animals received daily doses of lead acetate (4, 8, 12.5 mg kg−1 i.p.) for 28 days. GFAP concentrations were subsequently determined in different brain regions. Already at the lowest dose level GFAP levels were found to be significantly increased in several brain regions, notably in different parts of the cortex, hippocampus and striatum, while cerebellum and brain stem remained unaffected. Neurobehavioural assessment of animals, preceding the neurochemical analysis, also revealed significant alterations of neuromuscular function, excitability and spontaneous activity. Methylmercury has caused a number of poisonings in man (Marsh, 1985) where it appears to affect in particular both the central and peripheral nervous system. In an animal experiment, adult rats were subchronically dosed with methylmercury (0.75 or 2 mg kg−1) for 28 days. Neurobehavioural assessment indicated that grip strength was significantly impaired. Neurochemical analysis of GFAP in the central nervous system was performed in selected brain regions and, in addition, in various segments of the spinal cord. Increased GFAP levels were observed in the cerebrum only in the frontal cortex, also in brain stem and in spinal cord. A further detailed analysis of brain stem sub-structures showed significantly enhanced GFAP levels in pons and medulla oblongata but not in midbrain. In spinal cord GFAP concentration was increased in specific sections, e.g. in the cervical and lumbar segments but not in the thoracic segment. The results with these particular examples of heavy metals have indicated the unsuspected presence of regions in the central nervous system with astrogliosis, as determined in a biochemical GFAP assay. The neuronal damage involved has not yet been confirmed independently, e.g. by assessment of synaptophysin. The possibilities remain, therefore, that reactive astrocytosis by these heavy metals may be indirectly a result of breaching the integrity of the blood-brain barrier (Bressler and Goldstein, 1991) or of a direct toxic action on astroglial cells (Selvin Testa et al., 1990; Stark et al., 1992). Manganese is a well recognized industrial neurotoxin associated with neurologic effects after prolonged exposure in occupational settings (Katz, 1985). The clinical manifestations of manganism bear a large similarity to those of Parkinson’s disease (PD). The neurodegenerative disease PD is characterized by a selective loss of neurons in the basal ganglia.

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Experimental studies were done to investigate the vulnerability of the basal ganglia in manganism. Manganese (as Mn2+) was applied intrastriatally to rats and region-specific brain damage was assessed by 45Ca determining regional accumulation using quantitative autoradiographic procedures developed in this laboratory (Gramsbergen and Van der Sluijs-Gelling, 1993). It appeared that Mn2+ induced a timedependent 45Ca accumulation in most of the regions constituting the basal ganglia, e.g. striatum and several other structures such as globus pallidus, entopeduncular nucleus, several thalamic nuclei and substantia nigra (Sloot et al., 1994). As monoaminergic neurotransmission plays a predominant role in the basal ganglia, recent studies have indicated that concentrations of biogenic amines such as dopamine and its major metabolites, serotonin and noradrenaline were also reduced in striatum by Mn2+. The kinetics of this process indicated that concentrations of most monoamines and metabolites were temporarily reduced except for dopamine and metabolites in striatum that remained at a permanently reduced level (>90 days) (Sloot et al., 1994). In order to demonstrate the specificity of the manganese effects, several other compounds were studied, including ferrous ions (Fe2+). An equimolar dose of Fe2+, applied intrastriatally, produced, however, a much more extensive and widespread 45Ca accumulation throughout the basal ganglia and, in addition, in nucleus accumbens and cerebral cortex. Ferrous ions were also three times more potent than manganese ions in causing depletion of dopamine in striatum (Sloot et al., 1994). The results based on both 45Ca accumulation and biogenic amine levels are in concordance with the hypothesis that the basal ganglia, which are enriched in iron and iron-binding proteins, represent a selective target for manganese. The role and fate of endogenous iron in the brain, and the basal ganglia in particular, under toxic conditions including chronic manganese exposure, merits further investigations. Oxidative stress by free radical formation may play a role in these toxic events. Transition metals are known as strong promoters of reactive oxygen species. Especially iron, as the ferrous ion (Fe2+), has been found to react with hydrogen peroxide to form the hydroxyl radical in the so-called Fenton reaction (Halliwell et al., 1992; Aust et al., 1993). It is thought that this mechanism plays an important direct role in iron poisoning (Aust et al., 1993). In an indirect way oxidative stress by iron may be initiated by chemicals that are able to ‘liberate’ iron from stores such as ferritin, transferrin, haemoglobin, etc. Recently, evidence has been obtained that a number of chemicals, including the pesticides paraquat and diquat, may release iron from ferritin in vivo as well as in vitro (Aust et al., 1993), involving organic radical and superoxide formation. These observations are particularly relevant for the interpretation of the observed neurotoxic effects of iron described above. Dopamine is relatively

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easily subjected to a process of auto-oxidation. The decrease in dopamine levels by iron may have been caused by oxygen radicals. Circumstantial evidence suggests that a similar mechanism of action may underly manganese neurotoxicity. The decrease in dopamine in the basal ganglia by manganese is also thought to occur through catalysis of dopamine oxidation (LeBel and Bondy, 1991), possibly involving radical oxygen species. An open question is whether manganese might participate in the ‘iron release’ hypothesis (Sloot and Gramsbergen, 1994). Current efforts are being made to determine free radical formation by iron and manganese and to relate this to dopamine depletion. For this purpose rats are given salicylic acid (SA) and subsequently the SA hydroxylation products are measured in cerebrospinal fluid and brain tissues as indices of hydroxyl radical formation. Developmental neurotoxins (PCBs) Concern has been raised about the long term consequences of low level intake of polychlorinated biphenyls (PCBs) with respect to neurotoxicity as it relates to nervous tissue development and intellectual performance in the juvenile and adult stages. Several epidemiological studies with infants have shown a negative correlation between PCB levels in cord blood and cognitive functions and a positive correlation between PCB levels and altered neurological parameters such as hypotonia and hyporeflexia (Rogan et al., 1988; Jacobson et al., 1990). Experimental studies in various species including primates have also provided arguments for neurotoxic effects in offspring after perinatal exposure to PCBs (Tilson et al., 1990). In this laboratory evidence has recently been obtained to indicate dramatic reduction in sexual behaviour and reproduction in offspring that was perinatally exposed to PCBs (Smits-van Proojie et al., 1993). In order to investigate eventual structural alterations in the CNS, pregnant Wistar WU rats were exposed to Aroclor 1254 on days 10–16 of gestation. At a young age (3 weeks) and adult age (3 months) offspring were sacrificed and various brain regions were dissected for assessment of gliotypic and neurotypic proteins. In untreated control animals developmental aspects of astrocytes in the central nervous system were encountered. Both in hypothalamus and cerebellum of control animals GFAP levels were increased by 200–300 per cent between 3 weeks and 3 months postnatally. A developmental GFAP increase was also found in brain stem, striatum and lateral olfactory tract, although to a lesser extent. In hippocampus and prefrontal cortex, GFAP levels remained virtually unchanged between 3 weeks and 3 months. These results, therefore, indicate that glial cell maturation and/or differentiation is not uniformly distributed over the whole brain. It appeared that glial cells at birth in rats were fully developed in brain regions dealing with cognitive functions, e.g.

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cortex and hippocampus, while in regions such as hypothalamus and cerebellum this process occurred entirely neonatally. It should be kept in mind that in the rat also neuronal maturation and differentiation are not completed at birth and continue to a large extent postnatally. Exposure of pregnant rats to Aroclor 1254 did lead to a number of alterations in GFAP levels in various brain regions in offspring as compared to control animals. The most striking differences were observed in brain stem. While in control animals of both sexes GFAP levels increased postnatally, the normal developmental increase in perinatally exposed animals was absent (Figure 18.3, upper panel). At 3 months the relative deficit in GFAP levels was 41 per cent for male and 30 per cent for female progeny. Already at the lowest dose of Aroclor 1254 (5 mg/kg) a maximum decrease was observed (Figure 18.3, upper panel). In addition to brain stem, a similar effect on GFAP levels was found in striatum although the relative deficit at 3 months was somewhat less. Quite an opposite pattern emerged in brain regions such as cerebellum, lateral olfactory tract and prefrontal cortex, where GFAP levels were increased relative to unexposed progeny. In hippocampus no significant changes in GFAP levels were encountered. The question arose whether the observed neurodevelopmental toxicity in rats by PCBs was specific for astroglial cells or also involved neuronal maturation, differentiation and death. For this purpose the same nervous tissues were used for quantitative assessment of a neuronal marker in the form of synaptophysin. In brains of untreated adult control rats, regionspecific differences were observed in synaptophysin concentrations, being high in the prefrontal cortex, striatum and hippocampus and relatively low in lateral olfactory tract, cerebellum and brain stem. Synaptogenesis for the brain as a whole largely takes place in the rat from birth until postnatal day 70 (Knaus et al., 1986). Regional differences in the speed of postnatal synaptogenesis from 3 weeks to 3 months were found in a number of structures being most pronounced in cerebellum (190 per cent increase) and prefrontal cortex (170 per cent increase) while in brain stem little change was observed. As a result of perinatal exposure to Aroclor 1254, altered expression of synaptophysin was observed. In most brain structures examined, including brainstem (Figure 18.3, lower panel) significant decreases in synaptophysin concentrations were found. This suggests that during development of the central nervous system, PCBs may interfere with the formation of synaptic vesicles, synaptogenesis, or formation of nerve terminals. A straightforward interpretation of the present results is not possible at this stage. However, what is clear is that perinatal exposure to PCBs may cause changes in the structural composition of the central nervous system both in the neuronal and the glial cell compartment. Apparently there are different effects on nerve cells of the CNS depending on the brain region involved. The brain stem and striatum are regions with decreased

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Figure 18.3 Developmental effects on glial fibrillary acidic protein (GFAP) and synaptophysin concentrations in brain stem after perinatal exposure to Aroclor 1254. Results are expressed as mean±SEM with the values of the control animals at day 21 set at 100 per cent. Open bars, control; shaded bars, 5 mg Aroclor 1254 per kg; closed bars, 25 mg Aroclor 1254 per kg; D 21 and D 90, postnatal days; *, p<0. 05.

synaptophysin in association with decreased GFAP concentrations. One possibility is that in these structures postnatal maturation and differentiation of both neurons and glial cells may be halted or diverted. It may well be that this is a primary event in PCB developmental neurotoxicity. Neurodegeneration of developing and differentiating neurons is a possibility that might explain the combination of decreased

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Table 18.1 Cerebral calcium accumulation

a b

Systemic dose. Intrastriatal dose.

synaptophysin and increased GFAP in structures such as cerebellum, lateral olfactory tract, and prefrontal cortex. Such a situation could arise when neurons in these latter regions fail to receive proper input from developing neurons originating in brain stem and striatum. Further investigations into this interesting topic may provide further clues for the developmental toxicity of PCBs and related compounds that possibly could aid in the interpretation of neurobehavioural effects, for instance altered sexual behaviour and reproduction (Smits-van Prooije et al., 1993). The findings observed in cerebellum are consistent with a phase of hypothyroidism during development but it is clear that a conclusive role for thyroid hormone remains to be further established. The present results furthermore suggest that alterations in synaptophysin/GFAP levels may be useful and sensitive parameters to study compounds suspected of developmental neurotoxicity.

Conclusion The results with various neurotoxins demonstrate that assessment of gliotypic proteins such as GFAP may be a useful tool to identify and quantify persistent toxic insults of the CNS, especially when this is backed up by indications for loss of neuronal elements, e.g. decreased synaptophysin concentration. Also in circumstances of developmental neurotoxicity, caused by chemicals such as PCBs, an approach based on changes of gliotypic and neurotypic proteins may provide a promising biomarker. Of course, further investigations with various other compounds are required to substantiate these interesting findings. Cerebral calcium accumulation may be useful as an early indicator of neurotoxicity on a more prospective basis as is suggested by various examples of neurotoxic compounds (summarized in Table 18.1). Because in unlesioned brain regions the background levels of calcium uptake remain

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very low, areas where neurotoxic events take place are easily identified and quantified by autoradiographic or scintillation counting procedures. Finally, the role of oxidative stress by free radical formation in the nervous system merits further studies since it may open up possibilities for providing a biomarker that is linked to a mechanism of neurotoxic action. Acknowledgements The author is grateful for unpublished information provided by Dr Didema de Groot on neuropathology, by Dennis C.Morse, MSc and the students Wendelien Wesseling and Annemiek Plug, Agricultural University Wageningen, on developmental effects of PCBs; and for expert technical assistance by Alita van der Sluijs-Gelling. References ANDERSEN, M.E., 1991, Abstract, Eurotox Congress, Maastricht. AUST, S.D., CHIGNELL, C.F., BRAY, T.M., KALYANARAMAN, B. and MASON, R.P., 1993, Free radicals in toxicology, Toxicol. Appl. Pharmacol., 120, 168–78. BEN-ARI, Y., 1985, Limbic seizure and brain damage produced by kainic acid: mechanisms and relevance to human temporal lobe epilepsy, Neuroscience, 14, 375–403. BOOBIS, A .R., FAWTHROP, D.J. and DAVIES, D.S., 1989, Mechanisms of cell death, Trends Pharmacol. Sci., 10, 275–80. BORGHOFF, S., GARGAS, M.L., ANDERSEN, M.E. and CONOLLY, R.B., 1991, Development of a biologically based response model for 2,4,4-trimethyl-2pentanol induced alfa-2u-globulin nephropathy, Toxicologist, 11, 138. BRESSLER, J.P. and GOLDSTEIN, G.W., 1991, Mechanisms of lead neurotoxicity, Biochem. Pharmacol., 41, 479–84. BROCK, T.O. and O’CALLAGHAN, J.P., 1987, Quantitative changes in the synaptic vesicle proteins synapsin I and p38 and the astrocyte-specific protein glial fibrillary acidic protein are associated with chemical-induced injury to the rat central nervous system, J. Neurosci., 7, 931–42. CHERNIACK, M.G., 1988, Toxicological screening for organophosphorusinduced delayed neurotoxicity: complications in toxicity testing, Neurotoxicology, 9, 249– 72. DAVIS, C.S., JOHNSON, M.K. and RICHARDSON, R.J., 1985, Organophosphorus compounds, in O’Donoghue, J.L. (Ed.) Neurotoxicity of Industrial and Commercial Compounds, p. 1–24, Boca Raton, Florida: CRC Press. ENG, L.F., 1988, Regulation of glial intermediate filaments in astrogliosis, in Norenberg, M.D., Hertz, L. and Schoesboe, A. (Eds) The Biochemical Pathology of Astrocytes, pp. 79–90, New York: Alan R Liss. GRAHAM, D.G., ANTHONY, D.C., BOELKELHEIDE, K., MADSCHMANN, N.A., RICHARDS, R.G., WOLFRAM, J.W. and SHAW, B.R., 1982, Studies of

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SLOOT, W.N. and GRAMSBERGEN, J.B.P., 1994, Axonal transport of manganese and its relevance to selective neurotoxicity in the rat basal ganglia, Brain Res., 657, 124–32. SLOOT, W.N., VAN DER SLUIJS-GELLING, A.J. and GRAMSBERGEN, J.B.P., 1994, Selective lesions by manganese and extensive damage by iron after injection into rat striatum or hippocampus, J. Neurochem., 62, 205–16. SMITS-VAN PROOIJE, LAMMERS, J.H.C.M., WAALKENS-BERENDSEN, D.H., KULIG, B.M. and SNOEIJ, N.J., 1993, Effects of the PCB 3,4,5,3′,4′,5′hexachlorobiphenyl on the reproduction capacity of Wistar rats, Chemosphere, 27, 395– 400. SPENCER, P.S. and SCHAUMBERG, H.H., 1980, Experimental and Clinical Neurotoxicology, Baltimore: Williams and Wilkins. SPERK, G., LASSMAN, H., BARAN, H., KISH, S.J., SEITELBERGER, F. and HORNYKIEWICZ, O., 1983, Kainic acid induced seizures: neurochemical and histopathological changes, Neuroscience, 10, 1301–15. STARK, M., WOLFF, J.E. and KORBMACHER, A., 1992, Modulation of glial cell differentiation by exposure to lead and cadmium, Neurotoxicol. Teratol, 14, 247–52. TILSON, H.A., JACOBSON, J.L. and ROGAN, W.J., 1990, Polychlorinated biphenyls and the developing nervous system: cross-species comparisons, Neuro-toxicol. Teratol, 12, 239–48. VAN DEN BERG, K.J. and GRAMSBERGEN, J.B.P., 1993, Long-term changes in glial fibrillary acidic protein and calcium levels in rat hippocampus after a single systemic dose of kainic acid, Ann. N.Y. Acad. Sci., 679, 394–401. VERONESI, B., 1992, In vitro screening batteries for neurotoxicants, Neurotoxicology, 13, 185–96.

19 Endocrine Toxicology of the Thyroid for Industrial Compounds CHRISTOPHER K.ATTERWILL and SAMUEL P.AYLWARD CellTox Centre, University of Hertfordshire, Hatfield, Herts

General introduction Classification of endocrine toxicity Xenobiotic-induced endocrine dysfunction and toxicity is a common finding in safety studies and an increasingly important consideration in the riskassessment process. The effective identification of potential endocrine toxicological effects depends upon xenobiotic effect classification in relation to normal endocrine function and pathology. Classifications of different types of endocrine toxicity have been proposed in previous publications on this subject. Capen and Martin (1989) proposed a detailed classification based on clinical endocrine function and pathology. On the other hand, Baylis and Tunbridge (1985) proposed a simpler classification based on the adverse endocrine reactions of xenobiotics which are observed clinically. From a toxicological or preclinical safety testing point of view, Atterwill and Flack (1993) favoured classifying endocrine toxicology of xenobiotics in a manner which is similar in concept to that for classifying other toxicological phenomena (CIOMS, 1983) taking into account the unique nature of the endocrine system. This is as follows (see also Figure 19.1): Class 1 Effects which can be predicted from the endocrine pharmacology of compounds. An example would be the oestrogens and progestogens which have a plethora of effects on metabolic parameters in addition to their actions on oestrogen sensitive target sites when administered at pharmacological doses (the therapeutic dose levels). Further examples are certain dopamine and 5-HT antagonists acting on hypothalamic and pituitary receptors which may disrupt ‘downstream’ endocrine functions. Class 2 Effects which again can be predicted from the endocrine pharmacology of the compound when administered at doses well in excess of the therapeutic dose level. An example would be adrenal steroid suppression and general excessive catabolism observed with high dose and

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Figure 19.1 Classification of endocrine toxicity (Atterwill and Flack, 1993).

prolonged use of glucocorticoids. In addition, the use of thyromimetic agents which may suppress pituitary thyrotroph function. Class 3 Effects which could not have been predicted from the pharmacology of the compound. This group can be subclassified into: (a) Effects which are direct or primary actions on an endocrine gland. Examples of this might be the action of ketoconazoles on adrenal and testicular function and the action of alloxan and streptozocin on the β-cell of the pancreas; and (b) effects on endocrine glands which are indirect or secondary to changes in other organs or control mechanisms (homeostasis). Examples here would be the actions of phenobarbitone and PCBs on the rat thyroid and the effects of lactose and polyols on the rat adrenal medulla. Class 4 Effects which cannot be predicted from pre-clinical studies because of idiosyncratic effects on the endocrine system. As indicated by Baylis and Tunbridge (1985) the adverse effects of pharmaceutical agents on the endocrine system are generally due to normal

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or exaggerated pharmacological responses, that is Classes 1 and 2. Furthermore, it appears that endocrine toxicology can be detected reliably in pre-clinical studies. What is essential is that Class 3 toxic endocrine effects are classified appropriately. Class 3a effects would be expected to be seen across species, whereas there are numerous examples where Class 3b effects appear to be species-specific. There seems to be a paucity of clear-cut examples for Class 4 effects. There are several examples of compounds which have idiosyncratic immunotoxicological effects in susceptible humans which cannot be predicted from pre-clinical studies. Incidence of thyroid toxicity and tumours of the thyroid Most information on incidence is derived from pharmaceutical toxicity databases. Such data from Ribelin (1984) suggest that the endocrine system of the rats is particularly sensitive to toxicity from xenobiotics. This is also supported by Heywood (1984) in which he examined the target organ toxicity for 42 pharmaceutical compounds in the rat and dog. The endocrine system of the rat was only second to the liver as the most frequently affected target organ (38 per cent liver, 31 per cent endocrine). Ribelin (1984) reported that the most frequent endocrine lesion occurs in the adrenals followed by the testes but this analysis was conducted on chemicals and pharmaceuticals, and the data indicated that it was the cortical layers of the adrenal that were being predominantly effected, suggesting that the adrenal changes may be reflecting general stress responses rather than direct adrenal gland toxicity. In another analysis conducted in conjunction with the Centre of Medicines Research this area was further explored. Toxicology data on 124 compounds (all pharmaceuticals) were analysed. Just under 50 per cent (61/124) of these compounds have effects on one or more endocrine glands. Similar to Ribelin (1984) the adrenals were the most frequently affected, followed by the testes and the thyroid. An extensive survey of the different types of thyroid toxicity for both pharmaceutical agents and industrial chemicals was presented by Atterwill et al. (1993). Perturbation of thyroid function Thyroid function can be perturbed by agents affecting a number of processes involved in the regulation of the hypothalamic-pituitary-thyroidliver (H-P-T-L) axis (Figure 19.2). These agents can affect function directly by interacting with thyroid cell receptors on their intracellular transduction mechanisms (see Figure 19.3). Alternatively thyroid function may be altered indirectly by agents affecting thyroid hormone metabolism and/or distribution—this event being followed by the release of thyrotrophic factors, or by xenobiotic-mediated alterations in the release of these factors

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Figure 19.2 Hypothalamic-pituitary-thyroid-liver (HPTL) axis.

themselves from the hypothalamus or pituitary gland (see also review by Cavalieri and Pitt-Rivers (1981) and Atterwill et al. (1993)). For example, thyromimetics such as L-T3 or D-T3 can cause pituitary atrophy and thyroid ‘shutdown’ and toxicity due to suppressive effects on the thyrotrophs in the adenohypophysis (Atterwill, 1988).

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Figure 19.3 Thyroid follicular epithelial cell function.

Control of thyroid function and pathobiology of thyroid lesions Xenobiotic toxic effects on the hypothalamic-pituitarythyroid-liver (H-P-T-L) axis The control of mammalian thyroid follicular function is shown schematically in Figure 19.2 together with the points at which agents may perturb function and cause toxicity and thyroid lesions. The most common classes of agent affecting function are discussed at the different loci in terms of the categories (1–4) of endocrine toxicity (Atterwill and Flack, 1993). The various control mechanisms and factors influencing hormone synthesis, distribution and metabolism can be summarised as follows: thyroid hormone (T3 and T4) synthesis and secretion from thyroid gland are controlled by thyroid stimulating hormone (TSH) released from the pituitary gland. This in turn is under control by hypothalamic thyrotrophin releasing hormone (TRH) and circulating levels of the thyroid hormones. Thyroid hormones exist in the circulation in the free (free T4 (FT4) and free T3 (FT3)) and protein-bound forms (approx 99 per cent of total T4TT4 and total T3-TT3) and it is the FT3 hormone produced by deiodination from T4 which has both a physiological action on T3 nuclear receptors in target tissues and influences pituitary TSH output.

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Protein-binding of the hormone in the circulation takes place on several moieties, the profiles of which vary between different species (Dohler et al., 1973). The major proteins are thyroglobulin (TBG), thyroid binding pre albumin (TBPA) and albumin. Free and bound T3 and T4 are in dynamic equilibrium in the circulation and because of differences in affinity for the binding proteins there is more T3 in the unbound form (approx 0.4 per cent of total) than T4 (approx 0.04 per cent of total). Thyroid hormone synthesis takes place at the apical membrane of the polarised follicular epithelial cells and depends on TSH-stimulated active iodide uptake under the influence of many extracellular and intracellular factors and signals (see Figure 19.3). Iodide is oxidised by peroxidase enzymes which mediate the incorporation of iodine into the tyrosyl residues of the colloidal glycoprotein, thyroglobulin. Colloid-bound thyroid hormone is stored in the follicular lumen and is released into the circulation by lysosomal action on the colloidal complex following endocytosis of colloid droplets into the cells. The incorporation and organification of iodide into thyroid hormone within thyroid follicles and the toxicological effects of xenobiotics can be studied in vivo using the ‘perchlorate-discharge test’ (Atterwill et al., 1987) or in vitro using cultured thyrocytes (see Figure 19.4) from different species (Atterwill and Fowler, 1990). Catabolic metabolism of the hormonal products of the thyroid gland, FT3 and FT4, is achieved via two major pathways, deiodination and conjugation (glucuronidation and sulphation yielding a more water-soluble product for biliary excretion) and one minor route, deamination (decarboxylation, see Figure 19.9 later). The metabolic fate of thyroxine relies predominantly on its deiodination to T3, only 20 per cent of circulating T3 (the thyromimetic) is secreted by the thyroid (Engler and Burger, 1984). The remaining 80 per cent is derived from the deiodinative conversion of thyroxine (FT4) to T3. Deiodination can occur at several sites—the ones of major importance from the clearance aspect being liver and kidney whilst pituitary deiodination is essential for controlling responsiveness to circulating FT4 levels. The deiodinases exist as three iso-zymes: Type I (5′-D; localised in liver, kidney, thyroid and central nervous system (CNS) tissue; and is propylthiouracil (PTU) sensitive); Type II (5'-D; localised exclusively in the CNS, brown adipose tissue, and pituitary; PTU-insensitive); and Type III (5-D, CNS; PTU insensitive) and have different affinities for T4, different maturational patterns and different compensatory responses to hypothyroidism (see Kohrle et al., 1987). Bastomsky (1973), using Gunn rats, congenitally jaundiced due to UDPglucuronyl transferase deficiency (including T4-glucuronyltransferase) demonstrated that the rate limiting step in hepatic thyroxine clearance (via

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the biliary excretion pathway) is the formation of the glucuronic acid conjugate, T4-glucuronide. Hepatic conjugation, either sulphation (preferring FT3) or glucuronidation (preferring FT4) yields a more water soluble product, excreted in the bile/ biliary duct is a major pathway in T4 excretion. Further deiodination via the deiodinase group of enzymes of T4 T3conjugates, T4-amines and T3 to thyromimetically inactive iodothyronines (rT3 Triace, Tetrace, T2 and T1; see Figures 19.8 and 19.9) plays an important part in the T4/T3 biotransformation cascade and completes the thyroid hormone metabolic profile. The key factor in maintaining correct thyroid follicular capability is an appropriate TSH output to TRH stimulation alongside circulating levels of FT4. Perturbation of this homeostatic control results in a classical thyroid response. The initial thyroid responses to increasing TSH levels are follicular cell hypertrophy, loss of colloid and vascular dilatation. In conventional animal toxicology studies performed for regulatory authorities one of the first indices of thyrotoxicity, therefore, is the observation of altered thyroid histopathology, primarily as follicular cell hypertrophy and/or diffuse hyperplasia, often leading to focal hyperplasia, thyroid adenomas and adenocarcinomas in longer term toxicity studies after longer term exposure. Pathobiology of thyroid follicular cell hyperplasia and neoplasia Thyroid neoplasia (see Figure 19.5) develops predictably in experimental species exposed to any procedure inducing prolonged and excessive TSH secretion (for example, the administration of chemical goitrogens, chronic iodine deficiency or subtotal thyroidectomy) although humans and mouse appear to be more resistant to TSH-induced thyroid neoplasia than rat. The number of cytogenetic abnormalities within the thyroid epithelium increases with duration of excess TSH exposure, with follicular cell hyperplasia potentially leading to neoplasia. The histopathological sequence of events is as follows (see Zbinden, 1987): following hypertrophy of the follicular epithelial cells focal hyperplasias appear in the gland which are distinct areas of papillary growth with enlarged epithelia. As these foci continue to grow they form nodules partly surrounded by collagenous fibres. These lesions are transition states between focal hyperplasias and adenomas. Adenomas are larger nodules that compress the surrounding tissue and have a distinct capsule. Follicular microcarcinomas (characterised by irregular gland-like structures, basophilia and nuclear crowding) may appear in some nodules. Larger carcinomas usually retain a follicular structure but sometimes consist of solid sheets of polymorphous cells (Zbinden, 1987).

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Figure 19.4 Cultured porcine thyrocytes in vitro. These scanning electron micrographs show (a) individual cultured ‘inverted’ follicles, and (b) individual follicular epithelial cells at higher magnification (apical membrane facing upwards) displaying TSH-stimulated microvilli.

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Figure 19.5 Factors influencing the development of thyroid neoplasia.

Two consecutive processes are thought to occur in the development of thyroid tumours; first, initiation which occurs quickly and is irreversible and secondly, promotion which occurs slowly and is reversible (and for which cell proliferation may be a necessary but not sufficient condition). Initiators may be ionising radiation, chemical/biological agents or genetic factors, with TSH acting as a promotion agent. Spontaneous thyroid follicular cell tumours arise from unknown aetiologies and factors and ageinduced changes in the cell membrane, growth factor and signaltransduction mechanisms may be involved. Small subpopulations of hyperreactive epithelial cells retaining the high replication rate of the foetal stage have been identified (Peter et al., 1982; Smeds et al., 1987) which may enter clonal expansion following only slight elevations in TSH (such as those in handled or stressed animals) or other growth factors, leading to spontaneous nodular goitres (see Zbinden, 1987). Many experimental studies have confirmed the key role of TSH as a stimulator of thyroid growth. In a rat thyroid cell line (FRTL-5) TSHstimulated growth of the cells was found to be associated with a marked increase in c-fos and c-myc oncogene expression (Colletta et al., 1986). Another example of the tumour promoting capacity of TSH is given by

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studies where rats were given carcinogens such as N-methyl-N-nitrosourea (MNU) and then phenobarbital, or put on an iodine deficient diet. These treatments cause an early and increased incidence of thyroid follicular lesions and tumour formation (Hiasa et al., 1982; Oshima and Ward, 1984). The duration of exposure to high circulating TSH concentrations is also important in that intermittent administration of chemical goitrogens with TSH ‘normalisation’ does not appear to lead to follicular neoplasias. An elaborate series of studies have shown that a sustained elevation of serum TSH in the rat leads to three phases of thyroid growth (Figure 19.6): (1) a phase of rapid growth lasting 1–2 months, followed by (2) a plateau phase of 3–6 months (growth desensitising mechanism (GDM) limiting epithelial cell mitotic response), followed eventually (3) by the appearance of multiple follicular cell tumours (loss of GDM; see Wynford-Thomas et al., 1982; Stringer et al., 1985; Smith et al., 1986). The reversibility of TSHinduced thyroid focal hyperplasia will evidently depend, therefore, on the stage during these ‘timed’ cellular changes in the first 6 months at which the TSH stimulus is withdrawn. Once the GDM is non-operative reversibility is not possible. Tumour progression seems to occur by a multi-stage process involving clonal ‘expansion’ and naturally occuring clones of cells have been demonstrated with high intrinsic proliferation potential in the mouse thyroid gland (Smeds et al., 1987), perhaps helping to explain the focal nature of hyperplastic and neoplastic lesions. The loss of a GDM within the follicular cells appears to be accompanied by an altered dependence or sensitivity to certain growth factors as well as the possible loss of an antioncogene which limits the follicular cells’s growth response to TSH. For example, the growth of normal cultured human thyroid cells requires TSH and insulin-like growth factor 1 (IGF1) in combination, whereas cells from adenomatous tissue in vitro proliferate in response to either TSH or IGF independently (Williams et al., 1987). This is due to the acquisition of autocrine production of IGF1 by the tumour cells themselves (see Thomas and Williams, 1991). Since the differentiation and growth of thyrocytes under TSH is regulated by cyclic-AMP-dependent mechanisms (Figure 19.2), tissue hyperplasia and hyperthyroidism might be expected to result when activation of the adenyl cyclase-cAMP cascade becomes unregulated. This can occur, for example, when somatic mutations impair the GTPase activity of G-protein coupled reactors, which may thus behave as proto-oncogenes. Such a mechanism is probably responsible for the development of a minority of monoclonal hyperfunctioning thyroid adenomas (Parma et al., 1993) (these also result in a silencing of normal thyroid function in extra-adenomatous tissue). Other non-genotoxic factors, such as agents affecting patterns of DNA methylation when coupled with a growth stimulus, should also be given consideration when

Figure 19.6 Pathobiology of thyroid tumorigenesis.

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attempting to define mechanisms in thyroid carcinogenesis (Thomas and Williams, 1992). In summary, there is, therefore, good evidence that sustained TSH drive to the thyroid gland can lead to a de-regulation of thyroid function. When investigating xenobiotic or drug-induced thyroid tumour formation, the mechanisms whereby TSH drive is increased can be understood by undertaking a series of experimental studies using in vitro and in vivo techniques. Having delineated the mechanism of the thyrotoxic effect it may then be possible to determine whether a particular drug or compound elicits a similar response in different species (including humans) and to investigate the dose-response relationship for this effect. Investigative toxicological studies and examples of xenobiotics causing thyroid toxicity via the H-P-T-L axis Introduction Atterwill et al., (1993) give extensive examples of both pharmaceutical and industrial compounds causing thyroid toxicity via the five main sites along the H-P-T-L axis as shown in Figure 19.7 and readers should refer to this for further and more detailed information. In this chapter, three of these five thyroid toxicity loci are described in relation to the endocrine effects produced, industrial xenobiotic examples, and investigative in vivo and in vitro tests to delineate mechanisms and species-specific effects. This information is further summarised in Figure 19.8. In terms of industrial compounds the most frequently cited examples causing thyroid toxicity appear to be in the categories of: (i) those potentially affecting the plasma protein binding of thyroid hormones—for example, the nitrile herbicide, ioxynil (Ogilvie and Ramsden, 1988); (ii) those acting directly on the thyroidal peroxidase enzyme as goitrogens, and blocking thyroid hormone synthesis and secretion—for example, the coal derived hydroxyphenol products (Lindsay et al., 1992); and (iii) those affecting the hepatic metabolism and elimination T3 and T4—for example, compounds such as β-naphthoflavone, PCBs and alachlor (Ogilvie and Ramsden, 1988). Tables 19.1–19.3 show examples of these three class effects, compounds producing the effects and some of the range of investigative tests currently available. in vivo and in vitro studies of xenobiotics acting on the hepatic metabolism and clearance of thyroxine There is a growing list of agents, both pharmaceutical and industrial xenobiotics, which act in rodents by interfering with thyroid hormone

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Figure 19.7 Toxicological loci in H-P-T-L axis.

metabolism, hepatic elimination and thus circulating TSH levels (see also Capen and Martin, 1989; McClain, 1989; Atterwill et al., 1993). The xenobiotics include phenobarbital (McClain, 1989), β-naphthoflavone (Johnson et al., 1993), the polychlorinated biphenyls (Bastomsky, 1974), diproteverine (a calcium antagonist; Flack et al., 1989), SC37211 (a Searle imidazole antimicrobial (Comer et al., 1985), L649923 (a leukotriene D4 antagonist; Saunders et al., 1988), a novel oxyacetamide-FOE 5043 (Christenson et al., 1993), alachlor (Brewster et al., 1993), PCNB (pentachloronitrobenzene; Story et al., 1993), and hexachlorobenzene (Ogilvie and Ramsden, 1988). Most of these compounds have thus far been assumed to act in vivo via the induction of hepatic uridine diphosphate glucuronosyltransferase (UDPGT) in the rat, with species-specific formation of thyroid tumours in carcinogenicity studies (see McClain, 1989). Indeed many of the compounds, including phenobarbital do lead to increased hepatic UDP-GT activity and appearance of glucuronidated T4 in the bile, sometimes with elevated bile flow rates (see McClain, 1989). However, others such as the

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Figure 19.8 Investigative tests on H-P-T-L Axis.

pharmaceutical temelastine increase predominantly the clearance of free T4, though the bile product is not in conjugate form (Poole et al., 1989, 1990). Other compounds such as the food dye FD&C Red No 3 (Capen and Martin, 1989) are able to lower circulating triiodothyronine (T3) by altered deiodination suggesting the further existence of alternative mechanisms. Furthermore, not all chemicals inducing hepatic neoplasia in rodents cause thyroid neoplasia (McClain, 1989). We and others have reported that two SK&F histamine antagonists, temelastine (SK&F 93944) and lupitidine (SK&F 93479) produce ratspecific thyroid lesions via perturbation of the hepatic locus (Atterwill et

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Table 19.1

Table 19.2

al., 1989). Increased thyroxine clearance from the circulation, followed by elevated TSH ‘drive’ and increased thyroid follicular cell growth were observed. These com pounds act rapidly within minutes—hours of in vivo drug administration and are apparently able to increase the accumulation of T4 directly in vitro by cultured rat hepatocytes (Atterwill et al., 1989; Poole et al., 1990). Phenobarbital appears to share this property in vitro

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Table 19.3

(Aylward et al., 1994) and increases the accumulation of thyroxine in treated rat liver in vivo (Oppenheimer et al., 1968). Thyroxine transport (Figure 19.9) is regulated by specific components located within the plasma membrane in various cell types including fibroblasts and hepatocytes and is an important prerequisite for both hormone metabo lism and nuclear hormonally-mediated events (Pliam and Goldfine, 1977; Krenning et al., 1981; Blondeau, 1986). Investigations indicate that there exist two distinct transport systems specific to thyroxine: a high-affinity, low capacity, energy-dependent ATP-ase linked transport system and a low-affinity, high capacity transport mechanism (Sorimachi and Robbins, 1978; Krenning et al., 1981, 1983; Blondeau, 1986; Rao, 1991). Many of the compounds listed as indirect carcinogens in rat (due to an ability to induce liver microsomal enzymes and increase glucuronidated thyroxine elimination in the bile) have been shown to increase hepatic UDPGT activity or cause liver hypertrophy indicative of following repeated dosing (Comer et al., 1985; McClain, 1989; Johnson et al., 1993). However, there have been no attempts to provide a definitive link between UDP-GT induction and thyroid pathology or to prove a primary endocrinological effect via UDP-GT. Our previous work with temelastine in the rat in vivo (Atterwill et al., 1989) was able to demonstrate that the increased clearance of T4 from the rat circulation appeared within a few hours of a single compound dosing (Atterwill et al., 1989). Even

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Figure 19.9 Hepatic events leading to hormone elimination.

phenobarbital is able to increase the thyroxine clearance in a relatively short timespan (Atterwill, unpublished observations). New in vitro data are not inconsistent with the time course of the in vivo phenomena where enhanced T4 hepatocytic accumulation by cultured rat hepatocytes following compound exposure occurred as early as 60–90 min after exposure (Aylward et al., 1994). There was no membrane cytotoxic effect of the compounds at the threshold concentrations producing these effects in vitro. This shows a potential rapid direct effect of the xenobiotics on

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Table 19.4 Species and energy dependence of enhanced thyroxine accumulation in vitro

Key: ↑T4, increase; ↓T4, decrease; ↔T4, no change; Temperature? ATP?, temperature/ATP dependent; NA, not applicable.

hepatocellular T4 accumulation. The correlation between in vivo—in vitro species-specific toxicological effects is also evident (Figure 19.10 and Table 19.4). One of the features of temelastineinduced thyroid toxicology in vivo was the apparent species-specificity to the rat (Atterwill et al., 1989; Poole et al., 1989). Temelastine-mediated thyroid follicular hypertrophy and hyperplasia was not observed in dog, mouse or monkey following temelastine treatment (Figure 19.10). In vitro, no enhanced thyroxine accumulation in response to temelastine or phenobarbital was observed in guinea pig or dog hepatocytes (Aylward et al., 1994). In support of these findings, it has been demonstrated that the guinea pig is insensitive to thyroid pathological changes after phenobarbital or βnaphthoflavone administration in vivo (Johnson et al., 1993; Wyatt et al., 1993). In support of, and as an extension of these findings, we now present important new findings to demonstrate conclusively that some of the rapid ‘effectors’ of thyroid toxicity via the liver, such as temelastine, do so independently of a primary action on UDP-GT, whereas other cytochrome P450 inducers such as phenobarbital may have a combined effect. This work was carried out using hepatocytes prepared from UDP-GT system deficient Gunn rats. Studies on Gunn rat hepatocytes in vitro Hepatocytes were prepared from the normal or Gunn rat (deficient in UDPGT isozymes conjugating thyroxine) and exposed to either temelastine or phenobarbital (2 or 20 µM) for 3 h as before (Aylward et al., 1994). The results show (Figure 19.11) that whereas temelastine was able to enhance thyroxine accumulation in both types of hepatocytes, phenobarbital only produced alterations in hormone accumulation in normal cells, supporting

Figure 19.10 Effect of temelastine on 125I-T4 clearance (from Atterwill et al., (1989)).

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Figure 19.11 Effect of temelastine and phenobarbital on thyroxine accumulation in vitro by hepatocytes from control and Gunn rats.

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earlier in vivo findings where phenobarbital was toxicologically inactive in this strain of rat (Bastomsky, 1973).

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Conclusions The observations now lend further and strong support to the hypothesis that indirect xenobiotic-induced thyroid toxicology can arise from direct effects on hepatic membrane-located thyroxine transport proteins. It also suggests that the species-specificity of this toxic effect in vivo of some xenobiotics may be attributed to actual species differences in the sensitivity of these hepatic carriers to the compounds and not simply or primarily to changes in T4 glucuronidation via UDP-GT induction. For the first time we have demonstrated the usefulness of Gunn rat hepatocytes in vitro for discriminating between the two ‘hepatic subclasses’ of xenobiotics causing thyroid toxicity in rodents. A number of practical in vivo and in vitro investigative tests are now available for delineating mechanisms of thyroid toxicity along the H-P-T-L axis, and which also provide screening tools for examining chemical series of potentially toxic molecules: (i) Direct block of thyroid function via peroxidase inhibition can be measured in vivo by the perchlorate discharge test (Atterwill et al., 1987); (ii) it can also be measured in vitro using cultured thyrocytes (Atterwill and Fowler, 1990); (iii) indirect effects on hepatic thyroxine clearance can be assessed in vivo (Atterwill et al., 1989); or (iv) in vitro using cultured hepatocytes from different species or Gunn rat (Aylward et al., 1994). Effects on receptors at the hypothalamic and pituitary levels can also now be studied extensively using both in vivo and in vitro approaches (Buckingham and Gillies, 1993). This battery of technology now available will greatly advance the mechanistic understanding and screening of thyroid endocrine toxicants. References AKOSO, B.T., SLEIGHT, S.D., NACHREINER, R.F. and AUST, S.D., 1982, Effects of purified polybrominated biphenyl congeners on the thyroid and pituitary glands in rats, J. Am. College Toxicol., 3, 23–36. ATTERWILL, C.K. and FOWLER, K.F., 1990, A comparison of cultured rat FRTL-5 and porcine thyroid cells for predicting the thyroid toxicity of xenobiotics, Toxicol. in vitro, 4, 369–74. ATTERWILL, C.K. and FLACK, J.D., 1993, Endocrine Toxicology, Cambridge: Cambridge University Press. ATTERWILL, C.K., COLLINS, P., BROWN, G.G. and HARLAND, R.F., 1987, The perchlorate discharge test for examining thyroid function in rats, J. Pharmacol. Methods, 18, 199–203. ATTERWILL, C.K., KENNEDY, S., JONES, C.A., LEE, D.M., DAVIES, S. and POOLE, A., Comparison of the toxicity of orally administered Ltriiodothyronine (T3) in rat and cynomolgus monkey, paper presented at the British Toxicological Society Meeting, Oxford, March 1987.

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ATTERWILL, C.K., POOLE, A., JONES, R. and BROWN, R., 1989, Mechanistic investigation of species-specific thyroid lesions induced by treatment with the histamine H1 antagonist Temelastine (SK&F 93944) in rats, Food Chem. Toxicol., 27, 681–90. ATTERWILL, C.K., JONES, C. and BROWN, R., 1993, Thyroid Gland II— Mechanisms of species-dependent thyroid toxicity, hyperplasia and neoplasia induced by xenobiotics, in Atterwill, C.K. and Flack, J.D. (Eds) Endocrine Toxicology, pp. 137–82, Cambridge: Cambridge University Press. AYLWARD, S., WALKER, T. and ATTERWILL, C.K., 1994, Xenobiotic modulation of thyroxine uptake and efflux by cultured rat hepatocytes in vitro in relation to toxicological effects on the thyroid gland, Toxicol. in vitro (in press). BASTOMSKY, C.H., 1973, The biliary excretion of thyroxine and its glucuronic acid conjugate in normal and Gunn rats, Endocrinology, 92, 35–40. BASTOMSKY, C.H., 1974, Effects on a polychorinated biphenyl mixture (Arochlor 1254) and DDT on the biliary thyroxine excretion in rats, Endocrinology, 95, 1150–5. BAYLISS, P.H. and TUNNBRIDGE, W.M.G., 1985, Endocrine disorders, in Davies, D.M. (Ed.) Textbook of Adverse Drug Reactions, 3rd edn, pp. 335–51, Oxford: Oxford Medical Publications. BLONDEAU, J.P., 1986, Saturable binding of the thyroid hormone in isolated rat hepatocytes, FEBS Lett., 204, 41–7. BREWSTER, D.W., HOTZ, F., WARD, D.P., HEYDENS, W.E. and WILSON, A. G.E., 1993, Evidence for a hormonally-mediated non-genotoxic mechanism of action of alachlor-induced rat thyroid tumours, The Toxicologist, 13(1), 1440. BUCKINGHAM, J. and GILLIES, J.G., 1991, Hypothalamus and pituitary gland — xenobiotic-induced toxicity and models for its investigation, in Atterwill, C.K. and Flack, J.M. (Eds), Endocrine Toxicology, pp. 83–115, Cambridge: Cambridge University Press. CAPEN, C.C. and MARTIN, S.L., 1989, The effects of xenobiotics on the structure and function of the thyroid follicular and C-cells, Toxicolog. Pathol, 17(2), 266–92. CAVALIERI, R.R. and PITT-RIVERS, R., 1981, The effects of drugs on the distribu-tion and metabolism of thyroid hormones, Pharmacol. Rev., 33, 55– 80. CHRISTENSON, W.R., BECKER, B.D., HONNG, H.D. and WAHLE, B.S., 1993, Investigations into the mechanisms of decline in serum thyroxine following the administration of a novel oxyacetamide, (FOE 5043), The Toxicologist, 13(1), 1438. CI OMS, 1983, Safety requirements for the first use of new drugs and diagnostic agents in man. A review of safety issues in early clinical trials of drugs. Geneva: Council for International Organisations of Medical Sciences. COLLETTA, G., CIRAFICI, A.M. and VECCHIO, G., 1986, Induction of the c-fos oncogene by thyrotropic hormone in rat thyroid cells in culture, Science, 233, 458– 60. COMER, C.P., CHENGELIS, C.P., LEVIN, S. and KOTSONIS, F.M., 1985, Changes in thyroidal function and liver UDP-glucuronosyltransferase activity

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in rats following administration of a novel imidazole, (SC-37211), Toxicol. Appl. Pharmacol., 80, 427–318. DOCTER, R., KRENNING, E.P., BERNARD, H.F. and HENNEMANN, G., 1987, Active transport of iodothyronines into human cultured fibroblasts, J. Clin. Endocrinol. and Metabolism, 64(4), 624–8. DOHLER, K.D., WONG, C.C. and MUHLEN, A.V., 1973, The rat as a model for the study of drug effects on thyroid function: consideration of methodological problems, Pharmac. Ther., 5, 205–318. ENGLER, D. and BURGER, A.C., 1984, The deiodination of iodothyrones and their derivatives in man, Endocrin. Rev., 5, 151–2. FLACK, J.D., HAKANSSON, S., JEFFREY, D.J., KELVIN, A.S., MAILE, P.A., MCCURRDO, A.S. and PERKINS, C.I., 1989, Investigation of the effects of diproterivene on the thyroid of the rat, Hum. Toxicol., 8, 411. HEYWOOD, R., 1984, Prediction of adverse drug reactions from animal safety studies, in Bostrun, H. and Ljungstedt, N. (Eds) Detection and Prevention of Adverse Drug Reactions, pp. 177–89, Sweden: Almaqvuist & Wiskell Int. HIASA, Y., KITAHORI, Y., OSHIMA, M., FUJITA, T., YUASA, T., KONISHI, N. and MIYASHIRO, A., 1982, Promoting effects of phenobarbital and barbital on development of thyroid tumours in rats treated with N-bis (2hydroxypropyl) nitrosamine, Carcinogenesis, 3, 1187–90. JOHNSON, S., McKILLOP, D., MILLER, J. and SMITH, I.K., 1993, The effect on rat thyroid function of an hepatic microsomal enzyme inducer, Hum., Exp. Toxicol., 12, 153–8. JONES, C.A., BROWN, G.C., DICKENS, T.A. and ATTERWILL, C.K., 1988, Differential effects of D-a and l-isomers of triiodothyrmine on pituitary TSH secretion and peripheral deiodinase activity in the rat, Toxicology, 48, 273–84. KOHRLE, J., BRABANT, G. and HESCH, R.D., 1987, Metabolism of thyroid hormones, Hormone Res., 26, 58–78. KRENNING, E., DOCTER, R., BERNARD, B., VISSA, T. and HENNEMANN, G., 1981, Characteristics of active transport of thyroid hormones into rat hepatocytes, Biochim. Biophys. Acta., 676, 314–20. KRENNING, E., DOCTER, R., VISSER, T.J. and HENNEMANN, G., 1983, Plasma membrane transport of thyroid hormones: its possible pathophysiological significance, J. Endocrinol, 6, 59–65. LINDSAY, R.H., HILL, J.B., GAITAN, E., COOKSEY, R.C. and JOLLEY, R.L., 1992, Anti-thyroid effects of coal-derived pollutants, J Toxicol. Environm. Hlth., 37, 467–81. McCLAIN, R.M., 1989, The significance of microsomal enzyme induction and altered thyroid function in rats: implications for thyroid gland neoplasia, Toxicol. Pathol., 17(2), 294–306. MOSSMAN, T., 1983, Rapid colorimetric assay for cellular growth and survival: application to proliferation and cytotoxicity assay, J. Immunolog. Methods, 65, 55–63. OGILVIE, L.M. and RAMSDEN, D.B., 1988, Ioxynil and 3,5,3'-triiodothyronine: comparison of binding to human plasma proteins, Toxicol. Lett., 44, 281–7. OPPENHEIMER, J.H., BERNSTEIN, G. and SURKS, M.L., 1968, Increase thyroxine turnover and thyroidal function after stimulation of hepatocellular binding of thyroxine by phenobarbital, J. Clin. Invest., 47, 1399–406.

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OSHIMA, M. and WARD, J.M., 1984, Promotion of N-methyl-N-nitrosoureainduced thyroid tumours by iodine deficiency in F334/NCr rats, J. Nat. Cancer Inst., 73, 289–96. PARMA, J. et al., 1993, Somatic mutations in the thyrotropin receptor gene cause hyperfunctioning thyroid adenomas, Nature, 365, 649–51. PETER, H.J., STUDER, H., FORSTER, R. and GERBER, H., 1982, The pathogenesis of ‘hot’ and ‘cold’ follicles in multinodular goitres, J. Clin. Endocrinol. Metal., 55, 941–6. PLIAM, N.B. and GOLDFINE, I.D., 1977, High affinity thyroid hormone binding sites on purified rat liver plasma membranes, Biochem. Biophys. Res. Comm., 19, 166–72. POOLE, A., BETTON, G.R., SALMON, G., SUTTON, T. and ATTERWILL, C.K., 1989, Comparative toxicology of temelastine; a novel non-sedating H1 antagonist in the dog, rat and the monkey, Fundam. Appl. Toxicol, 14, 71–83. 14. POOLE, A., PRITCHARD, D., JONES, R.B., CATTO, L. and LEONARD, T., 1990, In vivo biliary excretion and in vitro cellular accumulation of thyroxine in rats or cultured rat hepatocytes treated with a novel histamine H1 receptor antagonist, Arch. Toxicol, 64, 474–81. RAO, C.G., 1991, Mode of entry of steroid and thyroid hormones into cells, Mol. Cell Endocrinol., 21, 97–108. RIBELIN, W.E., 1984, The effects of drugs and chemicals upon the structure of the adrenal gland, Fundam. Appl. Toxicol, 4, 105–19. SAUNDERS, J.E., EIGENBURG, D.A., BRACHT, L.E., WANG, W.R. and VAN ZWEITEN, M.J., 1988, Thyroid and liver trophic changes in rat secondary to microsomal enzyme induction caused by an experimental leukotrien antagonist, (L-649,923), Toxicol. Appl. Pharmacol., 5, 378–87. SMEDS, S., PETERS, H.J., JORTSO, E., GERBER, H. and STUDER, H., 1987, Naturally occurring clones of cells with high intrinsic proliferation potential within the follicular epithelium of mouse thyroids, Cancer Res., 47, 1646–51. SMITH, P., WYNFORD-THOMAS, D., STRINGER, B.M.J. and WILLIAMS, E.D., 1986, Growth factor control of rat thyroid follicular cell proliferation, Endocrinology, 119, 1439–45. SORIMACHI, K. and ROBBINS, O., 1978, Uptake and metabolism of thyroid hormones by cultured monkey hepatocarcinoma cells. Effects of potassium cyanide and dinitrophenol, Biochim. Biophys. Acta, 542, 515–26. STORY, D.I., CARDONA, R.A. and LENGEN, M.R., 1993, Effect of dietary PCNB on circulating levels of T3, T4 and TSH in rats, The Toxicologist, 13(1), 1446. STRINGER, B.M.J., WYNFORD-THOMAS, D. and WILLIAMS, E.D., 1985, In vitro evidence for an intracellular mechanism limiting the thyroid follicular cell growth response to thyrotropin, Endocrinology, 116, 611–15. THOMAS, G.A. and WILLIAMS, E.D., 1991, Evidence for and possiblemechanisms of non-genotoxic carcinogenesis in the rodent thyroid, Mutat. Res., 248, 357– 70. THOMAS, G.A. and WILLIAMS, E.D., 1992, Production of thyroid tumours in mice by demethylating agents, Carcinogenesis, 13, 1039–42.

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WILLIAMS, D.W., WYNFORD-THOMAS, D. and WILLIAMS, E.D., 1987, Human thyroid adenomas show escape from IGF-1 dependence for growth, Ann. Endocrinol., 48(2), 11. WYATT, I., GYTE, A., FOSTER, J.R., WILLIAMS, S.M. and ELCOMBE, C.R., 1993, Differences in the response of the rat and guinea pig liver and thyroid to hepatic enzyme inducers, Hum. Exp. Toxicol., 12(6), 561. WYNFORD-THOMAS, D., STRINGER, B.M. J. and WILLIAMS, E.D., 1982, Dissociation of growth and function in the rat thyroid during prolonged goitrogen administration, Acta Endocrinol., 101, 210–16. ZBINDEN, G. 1987, Assessment of hyperplastic and neoplastic lesions of the thyroid gland, TIPS (Dec., 1987), 8, 511–14.

20 Testing and Evaluation for Reproductive Toxicity ANTHONY K.PALMER Huntingdon Research Centre, Huntingdon

Introduction Most of the presentations at this meeting refer to high level, scientific investigations of one or two, highly important, high production volume chemicals, for which an adverse effect has been demonstrated. They are studies of characterisation, because they elaborate on known effects using a wealth of available information. But, how were the adverse effects of these few substances first discovered, what were the initial clues? Sadly, for many, the observation of adverse effects in humans was the trigger to intensive investigations, which is akin to ‘shutting the stable door after the horse has bolted’. This presentation is concerned with detecting effects of substances for which little or no information is available and, preferably, before they cause harm to humans. This requires a different kind of science, for which the main asset is the ability to predict, with reasonable accuracy, possible activity from minimal information. It requires wide experience and a balance between imagination and pragmatism. These attributes are especially important for toxicity to reproduction, which triggers instinctive reactions in even the coolest and most objective scientist. Identifying the cause of adverse effects on human reproduction has long been surrounded by controversy and uncertainty. In respect of the evaluation of substances for reproductive toxicity this state of affairs seems likely to persist for years to come. The main obstacle to any attempt to rationalise the situation is that any discussion on evaluation almost inevitably gravitates to the black hole of regulatory guidelines. All guidelines are flawed because science fact is compromised by bureaucracy and science fiction. For many reasons, but especially the unwillingness of any establishment to change the status quo, guidelines provide the worst starting point for developing a strategy for evaluation. Most guidelines are concerned only with methods for gathering specified information. On its own this information (on hazard) is insufficient and needs to be supplemented with other information, from other sources, to

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Table 20.1 Production volume triggers for industrial chemicals

Table 20.2 EC Annex VII and VIII toxicity tests for industrial chemicals

Notes: Tests for ecotoxicity are not included. Progression, by rote, from base set to level 2 would involve duplication of effort in several areas. The state of confusion regarding testing for toxicity to reproduction is portrayed by the failure to decide on requirements at base set level and the curious mixture of old and new terminology regarding tests.

predict whether humans might be affected. Most guidelines are a watershed in a broader spectrum of testing and assessment. They represent a point at which it may be decided that the only way to gather more information is to take the final step of exposing humans. Exposure of humans provides the

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Figure 20.1 Overlap of toxicity.

only certain way to determine whether reproduction would be affected, but we need to do the best we can before taking the chance. The paradox in this is that exposure of humans is facilitated by failure to demonstrate toxicity in animals. But, lack of activity, being negative, cannot be proven, only presumed. To make this presumption investigations must be extensive and comprehensive to convey reasonable assurance that failure was not due to deficiencies in methodology. There is an exception. For industrial chemicals, testing of all substances to these criteria would be a monumental task, therefore, less stringent testing is allowed according to production volume, which serves as an approximation for the extent of the population likely to be exposed (Table 20.1). It is a strategy based on population risk, the downside of which is an increased risk to the individual. The strategy takes advantage of the fact that, in a small population, the chance of identifying cause and effect is poor. At the base set level and level 1, as outlined by the EC, all tests are equivalent only to the voluntary preliminary studies conducted for

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medicines, agrochemicals and food additives. They may be sufficient to detect potent toxicity but not comprehensive enough to allow presumption of the absence of hazard, especially as the base set does not include tests for reproductive toxicity (Table 20.2). As production volume increases, more extensive testing should be undertaken. Often, this has been neglected, prompting the development of the OECD guidelines 421 and 422. These tests were intended to recover a situation that never should have arisen. A better approach? There is no question that evaluation for reproductive toxicity could be improved considerably. The question is whether industry and agencies are willing to do so. It would require a change of attitude in industry and agencies alike. Industry’s ‘passive avoidance’ of testing would need to be replaced by ‘active participation’. For a new substance the first step should be an integrated assessment of commercial prospects and potential toxicity over a broad spectrum (Figure 20.1). Early identification of ‘serious bad actors’, which tend to effect many systems, can save time and effort. Given the prognosis of problems ahead, it may be better to devote resources to finding safer alternatives, or to risk management, rather than to endless testing. For materials with a high commercial potential, the aim should be to get to full scale tests by the quickest route. Following the EC levels by rote is very inefficient since there is duplication with successive steps. Methods With an active participation policy, a much broader scope of methodology can and should be considered, ranging from searches for structure-activity relationships, through various in vitro methods, whole animal tests, wild life surveys and human surveys (Table 20.3). Due to time constraints I will concentrate on whole animal test systems. Structure-activity databases Structure and activity relationships are an obvious place to start any evaluation, despite the fact that currently available databases are far from perfect. Their reliability could be improved dramatically by adding unused information currently hidden in industry and agency archives. Entire mammalian tests Tests in entire mammals provide the only way of assessing what effects a substance may evoke in the complex, integrated and dynamic process of

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Table 20.3 Methods for detecting effects on human reproduction

reproduction. To detect the wide range of possible effects it is necessary to expose mammals to a substance from conception through sexual maturity. It is necessary to look for consequences of this exposure through at least one life cycle (Figure 20.2). This long observation period is required for detection of latent manifestations of developmental toxicity, such as those induced by lead, alcohol, diethylstilboestrol and other hormonally active substances. The only means of covering all these aspects is a two generation study or the equivalent in a combination of tests. Restricted test systems The use of lesser tests for industrial chemicals is a concession. Examination for some effects is omitted because they are not perceived to be important or because they would be difficult to detect, or because they occur very rarely. For example, first detection of effects in offspring of second generations is rare so such activity does not have a high priority. With these restricted tests, emphasis should be on detecting effects and not on manipulating a no effect level. Tests that could be considered would include OECD 421 and 422, the OECD single generation study and the old FDA Segment I study for medicines. The latter two are restricted tests because they do not allow detection of latent manifestations of developmental toxicity. The best return for effort is afforded by OECD 422, which combines examination for general or systemic toxicity, as well as reproductive toxicity. However, realising its potential requires an experienced laboratory team, the courage to modify the test and the conceptual ability to know how to interpret the results.

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Figure 20.2 Cycle of life/reproduction.

OECD 421 involves treatment of both sexes from about 2 weeks prior to mating through to termination, a few days after birth of offspring (Figure 20.3). Assessment of male fertility is achieved in two parts. Males are paired with females for detection of effects unrelated to spermatogenesis, for example, effects on sexual behaviour, libido or ejaculation and functional maturation of sperm. For detecting effects on spermatogenesis, direct methods, particularly histopathological examinations of testes and epididymides are used. Sperm analysis (seminology) could be added, although it does not seem to be better than histopathology. These methods could be incorporated into systemic toxicity studies rather than in the reproduction study per se. In respect of fecundity, treatment and observations of females include most of those applied in full scale tests. An exception is the lack of observations for delayed, post-natal manifestations. Detailed examination of foetuses for skeletal and soft tissue abnormalities is not included. The potential for prenatal effects is deduced by observation of post-natal differences in numbers pregnant, litter size, litter and mean pup weight at birth and to day 4 post partum. As with any guideline, OECD 421 should be used with commonsense and flexibility. If pretesting prognosis suggests that prenatal effects are unlikely, extension of the study to weaning of the offspring (Figure 20.3) provides added safeguards at little extra cost. Increase group size and it

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Figure 20.3 OECD 421 priority selection test.

Figure 20.4 Fertility and embryotoxicity.

provides the equivalent of the OECD single generation study. Conversely, if pretesting prognosis suggests a high probability of prenatal effects, including induction of malformation, then females could be killed just before delivery and foetuses examined for structural defects (Figure 20.4). This provides the equivalent of a fertility and embryotoxicity study we will see again later. OECD guideline 422 simply adds to OECD 421, elements for assessment of systemic and neurotoxicity. For those who have never conducted such a test it seems impossibly complex, but it is neither as difficult to perform, nor to interpret, as is feared. Its rejection by EC Officialdom makes it an even better proposition, since there need be no inhibitions about modifying the design according to circumstances. Some brief examples of results that may be encountered with positive materials are illustrated by the examples of Carbendazim (metabolite of Benomyl), DEHP, Cyclophosphamide and ethylene glycol methyl ether (EGME, 2-methoxyethanol). With Carbendazim (Table 20.4) macroscopic and microscopic examinations show unequivocal effects on testes and epididymides indicating an effect on spermatogenesis. An effect on females and offspring is indicated by an increased duration of pregnancy, reduction in the number of females with live young and lower values for litter size,

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Table 20.4 OECD 422: carbendazim, tabular summary

Notes: a Malformations included hydrocephaly and misaligned tails. pp=post partum Bold type indicates treatment effects including macroscopic and microscopic changes in testes and epididymides (an effect on spermatogenesis), an increased duration of pregnancy, reduction in the number of females with live young and lower values for litter size, litter weight and mean pup weight. The dosage related pattern of response provides added emphasis, as does the observation of malformed foetuses.

litter weight and mean pup weight. The dosage related pattern of response provides added emphasis, as does the observation of malformed foetuses. With DEHP (Table 20.5) an effect on spermatogenesis is evident at 2000 mg kg−1. Treatment at this dosage had to be withdrawn shortly after mating to avoid further mortalities of the more susceptible females. There is a marked reduction in the number of pregnancies. At lower dosages an increased duration of pregnancy, reduced litter size and litter weight, indicates an effect on the female and/or offspring. The higher mean pup weights are consequent to the longer duration of pregnancy. With cyclophosphamide (Table 20.6) female deaths at 3 and 4.5 mg kg−1 are attributable to treatment as, at 6.7 mg kg−1, all females died. There was a reduction in the number of females with young, litter size, litter weight and mean pup weight, providing clear evidence of effects on the female and

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Table 20.5 OECD 422: DEHP, tabular summary

Notes: [] Treatment at 2000 mg kg−1 was withdrawn after mating (4 weeks of treatment) due to loss of condition and mortality of females. pp=post partum. Bold type indicates treatment effects on spermatogenesis at 2000 mg kg−1. There is a marked reduction in the number of pregnancies. At lower dosages an increased duration of pregnancy, reduced litter size and litter weight, indicates an effect on the female and or offspring. The higher mean pup weights are consequent to the longer duration of pregnancy.

offspring. No effects on spermatogenesis were reported but, if a dosage inducing effects on the male had been selected, all the females would have died. With EGME (Table 20.7), dosages were based on acute toxicity and limited repeat dose toxicity studies only. This provided a more realistic representation of the testing of a new substance. The first consequence was the occurrence of systemic toxicity at 500 and 1000 mg kg−1. Treatment at the high dosages had to be withdrawn prior to mating, investigating recovery became a new objective. At 100 mg kg−1 profound effects on spermatogenesis were evident. Some pregnancies were obtained, but no live young were born. For the high dosage groups, effects on males remained evident several weeks after withdrawal of treatment. The duration of pregnancy was increased. A consequence of this is the higher mean pup weight. Values for numbers of implantations, live young and litter weight were lower than control values. So, not only did the test show the anticipated effects on reproduction, it also demonstrated that recovery of

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Table 20.6 OECD 422: cyclophosphamide, tabular summary

Notes: pp=post partum. Bold type indicates treatment effect. Female deaths at 3 and 4.5 mg kg−1 are attributable to treatment as, at 6.7 mg kg−1, all females died. There was a reduction in the number of females with young, litter size, litter weight and mean pup weight. No effects on spermatogenesis were reported but, if a dosage inducing effect on the male had been selected, all the females would have died.

females was slow. As far as I know this has not been mentioned in the extensive literature on EGME. These examples and others show that the OECD tests are capable of detecting substances with marked effects on reproduction. With such results it would be foolhardy to consider higher level guideline tests. If the substance is not abandoned any further testing would require case by case designs to characterise the detected effects more completely. Full scale testing If pretesting prognosis suggests that a substance is unlikely to present problems of toxicity, fast track progression to level 2 testing should be considered (Table 20.2) to avoid unnecessary duplication of step by step testing. Expected tests include a two generation study in rats and an embryotoxicity study in rats and rabbits. In the harmonised guideline for medicines this same strategy is just one of several options and the same or greater flexibility should be made available for testing industrial chemicals. Tests for embryotoxicity An anomaly in this strategy is the specification for embryotoxicity studies in two species, when only one species is required for all other aspects of

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Table 20.7 OECD 422: EGME, tabular summary

Notes: [] Treatment at 500 and 1000 mg kg−1 withdrawn prior to mating due to loss of condition and mortality. Animals in withdrawal phase. NE not examined pp=post partum Bold type indicates treatment effect. At 100 mg kg−1 profound effects on spermatogenesis were evident. Some pregnancies were obtained, but no live young were born indicating effects on females and the conceptus. For the high dosage groups, effects on testes and epididymides remained evident several weeks after withdrawal of treatment. The duration of pregnancy was increased with a consequent increase in mean pup weight. Values for numbers of implantations, live young and litter weight were reduced indicating slow recovery. This does not appear to have been mentioned in the extensive literature on EGME.

reproductive toxicity. A more sensible strategy would be to identify a relevant species before testing. It is pointless to conduct a test in an unsuitable species and doubly pointless to conduct tests in two irrelevant species. The requirement for detailed examination of foetuses for abnormalities is based on an exaggerated perception of risk prompted by fear. For many reasons the risk of inducing abnormalities is extremely low. Those same reasons make detection by direct observation of malformations unreliable.

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Table 20.8 Post-natal detection of prenatal effects

Notes: Prenatal effects include any soft tissue or skeletal changes (variants, anomalies or malformations) or altered foetal weight. Post-natal effects include reduced litter size at birth, reduced mean foetal weight or increased post natal mortality. In the few studies in which post-natal effects were not detected, prenatal changes were of a ‘minor’ nature (e.g. reduced ossification) and sometimes not conclusive. The lower number of litters reared in the Japanese Experiment 2 design would contribute to the slightly higher rate of ‘failures’. Where more serious prenatal effects such as the observation of malformations or prenatal death were observed a post-natal effect was always observed.

The dimensions of the tests we conduct are too small and dosage regimes are contradictory to the basic principle that malformations are induced by application of a precise dosage at a precise time. Whilst direct observation of malformations is unreliable the saving grace is that induced malformations always occur within a wider spectrum of embryotoxicity. This wider spectrum provides a more reliable, if indirect, means of detecting substances that might cause malformations. Also, these effects are important in their own right. Major effects such as altered offspring weight and prenatal death can be observed postnatally. For example both EC Segment I and Japanese Experiment 2 studies include examinations for foetal abnormalities as well as postnatal observations. A survey of such studies shows that, when malformations were observed, a post-natal effect was always observed (Table 20.8). Surveys such as this should have been conducted or sponsored by agencies and industry before formulating guidelines. Why have they not done so? The obsession with abnormalities is based on the fear of another thalidomide. The great contradiction is that a considerable number of rat embryotoxicity studies with thalidomide failed to provided convincing demonstration of teratogenicity. Conversely, reproduction studies in rats provided unequivocal, post-natal evidence of effect in the form of a marked reduction in the number of females with young and a marked reduction in

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Table 20.9 Thalidomide—rat reproduction studies

Notes: Thalidomide was administered in the diet to provide a daily intake of 200 mg kg−1 bodyweight from 60 days prior to mating (US FDA two litter test). Later studies demonstrated that the reduced percentage of females with young and the lower live litter size in females with young was associated with embryolethality.

litter size of the few that had young. Two studies, each containing several matings showed the reproducibility of the results (Table 20.9). Two generation studies The emphasis on structural abnormalities detracts from examination for other, important and more likely manifestations of reproductive toxicity (Figure 20.5). For example, current and proposed test guidelines for nonmedicines lack procedures for detection of developmental neurotoxicity (or behaviour). This is a curious contradiction given the current fashion for investigation of adult neurotoxicity. For detecting other effects on reproduction, all regulatory versions of the two generation study leave something to be desired. Even more disappointing is that newer versions proposed by the US FDA, the US EPA, the EC and OECD have recycled many of the old flaws. Truly, there has been a great deal of activity but very little progress. All current and proposed guidelines continue to require a prolonged premating treatment period for the F0 or parent generation. The claim that it is necessary to treat males for a full spermatogenic cycle is pure science fiction. Spermatogenesis is not a cycle but a sequence of overlapping batch processes. At any one time, all stages of spermatogenesis are present and a short dosing period is sufficient to cause effects. To detect these effects, direct histopathological methods can be applied shortly after treatment. Direct methods are quicker and more certain than mating to females. Mating trials are inefficient and lack sensitivity due to

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Figure 20.5 Manifestations of developmental toxicity

the high sperm production capacity of animals compared with humans. Interim results from an ongoing survey show that, where data are available, direct methods are effective and that a combination of direct methods with a premating dosing time of 2 weeks or less is even more effective (Table 20.10). The combination compares favourably with prolonged premating treatment and mating. Why have agencies and industry failed to conduct or sponsor such surveys? For non-medicines, use of a prolonged premating treatment period for the parent generation is an unnecessary duplication as treatment is continued into the F1 generation; this cannot be mated until animals have reached sexual maturity. Using science facts, a more efficient design for a two generation study (Figure 20.6) would include the following features: – One control and 2–4 test groups with dosages set at 2–5 fold descending intervals from a high dosage. – The high dosage should be a limit dose (1000 mg kg−1) or one inducing a minimal systemic effect on adults. – A short 2–4 week premating treatment period for both sexes of the parent (F0) generation. – A greater group size for the F0 generation to allow balanced selection of the F1 generation.

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Table 20.10 Detecting effects on males

Notes: a Unknown=data not available or not examined A, organ weights, histopathology, serum chemistry. B, sperm analysis, count, motility, morphology. Data derived from an ongoing survey of 150 substances for which an effect on males has been claimed, mostly from human studies. To date 80 of the substances have been evaluated. Results indicate that use of a prolonged premating treatment period (>2 wks) has not been helpful for indicating effects on humans and is no better than use of a short premating period alone (<2 wks). Results under the heading ‘detectable’ would indicate that a short premating treatment period of 2 weeks combined with a direct method would provide the most effective means of detection. Such a combination was advocated in all drafts of the ICH guideline, except the final draft 17 which recommended a 4 week premating treatment period for males. No evidence was provided to indicate that a 4 week premating treatment period of males is better than 2 weeks.

– Decision points for introduction of result based options such as a second mating or histopathological examinations. Results may be those within the study or from parallel studies. – Inclusion of simple procedures for detection of developmental neurotoxicity. Alternatives to the two generation study Any two generation study is best used only when there seems little possibility of an effect on reproduction. It is a test for the unexpected. With very active substances, it may be difficult to select dosages providing reassuring margins above potential human exposures. The need to cover an entire life cycle would make it time consuming for a non-rodent species. In such cases study combinations adding up to the equivalent of a two generation study could be used (Figure 20.7). It is not difficult to devise a sequence from a single study through two and three study combinations, to case by case designs. For the wide variety of industrial chemicals, case by case design should be the rule rather than the exception.

Figure 20.6 A two generation study.

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For example, a simple division of a two generation study will provide a developmental toxicity study in which pregnant females are treated from implantation through lactation or beyond and an F1 generation reared through to sexual maturity (Figure 20.8). The counterpart to this is a study of fertility in which both sexes are treated from about 2 weeks prior to

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Figure 20.7 Reproductive toxicity—selecting studies.

Figure 20.8 Pre- and post-natal (developmental toxicity) study.

mating through to termination of males after a minimum of 4 weeks of treatment overall (Figure 20.9). Treatment of females continues to implantation and they may be killed and examined at about day 13–15 of pregnancy. Alternatively, treatment of females can be continued beyond closure of the palate or even through pregnancy (Figure 20.4). Foetuses can be

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Figure 20.9 Fertility

delivered and examined for abnormalities according to procedures used in embryotoxicity studies. (Note that this study is almost identical to a modified OECD 421 study.) Combination with the ICH developmental toxicity study provides the equivalent of a two generation and embryotoxicity study. Interpretation of studies For regulatory agencies, one of the purposes of these tests is to gather information for labelling. By common consensus, a substance is labelled as a reproductive toxicant only if it induces effects at dosages below those causing systemic toxicity. Such labelling, however, can be very misleading especially with industrial chemicals. Even if the animal species can be shown to be a good surrogate for humans by means of kinetic and other studies, it is necessary to take into account the relationship between exposures causing effects and those likely to be encountered by humans. For example, a material can be labelled as a reproductive toxicant but present little or no real risk to humans because the effects are induced at exposures well in excess of those encountered by humans (Figure 20.10). Conversely, a substance not labelled as a reproductive toxicant could cause reproductive effects in humans if these are induced at exposures encountered by humans. In other words toxicity is relative, a matter of dosage and situation. Labelling without reference to exposures is incomplete.

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Figure 20.10 Interpretation/extrapolation of reproductive toxicology.

Conclusions In conclusion I should mention that time constraints enforce superficial mention of many important aspects. I would like to emphasise that improved testing and evaluation for toxicity to reproduction can be achieved with methodology that exists within and without regulatory guidelines. This has been illustrated by the special cases presented at this meeting. There is no necessity to be restricted to specific guidelines, there never was. We should make use of any and all test methods available as appropriate for the substance being investigated. Whatever the type of substance, testing involves looking for the same hazards. Having identified a hazard, methods of assessing risk, essentially, are the same (although PBPK models for reproductive toxicity would be more complex than those used for systemic toxicity). The methodology is available, what is required is the willingness and wisdom to use it effectively and efficiently (Palmer, 1993a, b). The goal should be to investigate a specific substance to the extent necessary, no more and no less. For this there needs to be a change in attitude by industry, agencies and academia. Therein is the greatest problem since ‘Change is not made without inconvenience, even from worse to better’ and people are very unwilling to change their prejudices and habits.

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References PALMER, A.K., 1993a, Identifying environmental factors harmful to reproduction, Environm. Hlth Perspect. Supplements, 101(2), 19–25. PALMER, A.K., 1993b, Introduction to (pre)screening methods, Reproduct. Toxicol., 7, 95–8.

PART SIX Toxicity of selected classes of industrial chemicals

21 Special Points in the Toxicity Assessment of Colorants (Dyes and Pigments) HERMANN M.BOLT Institut für Arbeitsphysiologie an der Universität Dortmund, Dortmund

Introduction Colorants (dyes and pigments) are very important industrial chemicals. A special point in the toxicological assessment of such compounds is their bioavailability upon inhalation. From the technological point of view pigments are colorants which are insoluble whereas dyes are soluble in the application mixture. Biologically, the most relevant route of potential exposure of humans to colorants is by inhalation. If a pigment is biologically insoluble, it may finally be removed from the airways by clearance mechanisms. However, in practice the situation is much more complicated. For instance, chromates are technically important pigments which are well investigated. Biochemical and toxicological research has shown that the common toxicological principle of chromates which penetrate the cell membrane and, after intracellular transformation, exert genotoxic effects, is the chromate anion (CrO ). In terms of inhalatory carcinogenicity, the very water-soluble alkali chromates and the practically insoluble lead chromate have the lowest potency. Pigments of an intermediate solubility, e.g. calcium chromate, zinc chromate, strontium chromate, have a high carcinogenic potency on the respiratory tract. Local storage of chromate particles in the airways with a slow but continuous local release of CrO seems therefore to be an important factor in respiratory carcinogenesis induced by chromates. But also lead chromate, which is technically regarded as insoluble, is bioavailable to some extent; this is visualized by practical cases of occupational lead chromate exposure which display markedly elevated blood lead levels. The question of systemic bioavailability, upon inhalation, became of recent regulatory importance for azo colorants based on carcinogenic aromatic amines. This problem has already been addressed in detail (Myslak and Bolt, 1988).

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Table 21.1 Number of azo colorants based on cancerogenic aromatic amines (2napthylamine, benzidine and its derivatives) listed in Colour Index (3rd edn, 3rd Rev., 1987)

The problem of carcinogenic azo colorants In the past, azo colorants based on benzidene, 3,3′-dichlorobenzidine, 3,3′dimethylbenzidine (o-tolidine), and 3,3'-dimethoxybenzidine (odianisidine) have been synthesized in large amounts and numbers, especially in the German chemical industry. The Colour Index (1987) lists a total number of more than 2000 azodyes, 452 of them being based on 2naphthylamine, benzidine, or benzidine derivatives (Table 21.1). Azo colorants have a number of properties that have made them invaluable for dyeing a wide variety of materials, including natural and artificial fibres, plastics, resins, textiles, leather, paper, glass, ceramics, cement, inks, printing inks, chalks, crayons and carbon papers, as well as cosmetics, food and beverages. Interesting with respect to potential exposure of painters is the use of azo colorants in the coloring of oil-, resins-, emulsion-, lime-, and other aqueousbased paints, distempers, transparent laquers, spirit and oil wood stains, and varnishes (Colour Index, 1987). In all these fields, particularly benzidine-based azo colorants have found widespread use (Gregory, 1984). In the UK, the Carcinogenic Substances Regulation led in 1967 to discontinuation of the use of benzidine in the production of azo colorants (Martin and Kennelly, 1985). The US government in 1974 promulgated regulations to control benzidine at the workplace (Gregory, 1984). Nevertheless, in the period of 1972–4, more than 150000 persons in the USA were potentially occupationally exposed to benzidine-based colorants (Gregory, 1984); in 1978, approximately 1.7 million US pounds of benzidine-based azo colorants were manufactured, and a further 1.6 million pounds were imported into the USA (Lynn et al., 1980). In Germany, over 30 different benzidine-based azo colorants were manufactured in the early 1960s. The manufacture of these colorants was stopped in 1971, with the exception of one dye (Direct Black 4; C.I. No. 30245); the manufacture of the latter was continued until 1973. Azo

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colorants based on carcinogenic congeners of benzidine (e.g. 3,3′dimethoxybenzidine; 3,3′-dimethylbenzidine) are most likely still being manufactured in some countries. The case of pigments based on 3,3′dichlorobenzidine is discussed below. Azo colorants are biologically active through their metabolites. Azoreduction of these compounds occurs in vivo (Radomski and Mellinger, 1962; Rinde and Troll, 1975; Robens et al., 1980) by an enzyme-mediated reaction. Azoreductases are found in mammalian tissues, particularly in liver (Fouts et al., 1957; Walker, 1970; Martin and Kennelly, 1981; Kennelly et al., 1982) and also in gut bacteria (Yoshida and Miyakawa, 1973; Chung et al., 1978; Hartmann et al., 1978; Cerniglia et al., 1982; Bos et al., 1986). The result of this azoreduction is the release of the (carcinogenic) aromatic amine from the colorant (Martin and Kennelly, 1985). Studies performed on exposed workers have demonstrated that the azoreduction of benzidine-based colorants occurs in man (Genin, 1977; Boeninger, 1978; Lowry et al., 1980; Meal et al., 1981; Dewan et al., 1988). Studies of Lynn et al., (1980) and Bowman et al. (1983) have demonstrated that the metabolic conversion of benzidine-, 3,3′dimethylbenzidine- and 3,3′-dimethoxybenzidine-based colorants to their (carcinogenic) amine precursors in vivo is a general phenomenon that must be expected for each member of this class of chemicals. However, in contrast to water-soluble dyes, the question of biological azoreduction of (practically insoluble) pigments was a matter of discussion in the recent years. One study has claimed the presence of 3,3'dichlorobenzidine in the urine both of experimental animals fed with Pigment Yellow 12 and of exposed workers (Akiyama, 1970). However, other experimental studies, using more modern analytical tools, did not confirm these results (DHEW, 1978; Leuschner, 1978; Mondino et al., 1978; Nony et al., 1980). Several epidemiological studies have demonstrated that the use of the benzidine-based dyes has caused bladder cancer in humans. In a Japanese study, the risk of bladder cancer among dye applicators (kimono painters) was 6.8 times the expected rate (Yoshida et al., 1971). In a British study, workers performing the dyeing of textiles (and not exposed to benzidine itself) had a higher risk of bladder cancer (RR=3.4) than expected (Anthony, 1974). In a USSR study, an increased incidence of bladder cancer was found in workers who dried or ground benzidine-based dyes (Genin, 1977). In our own study on bladder cancer in painters (Myslak et al., 1991), the time of first exposure (painters with bladder tumors) dated mostly back to the first half of the century. Two factors may have been relevant: (1) at that time, a large number of benzidine-based azodyes was in manufacture, especially in Germany; (2) during that time it was usual for painters in Germany to prepare their paints themselves. This work included grinding

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and mixing of the dyes and preparation of the coloring mixture by addition of solvents. All painters we had interviewed reported that this type of work had been regularly associated with considerable occurrence of dye dust in the atmosphere, up to the end of the 1950s. The very long latency period may explain why an enhanced risk of bladder cancer in German painters (due to previous exposure to azo dyes) is observed even today. Similar arguments have also been put forward for other occupational groups associated with an increased risk of bladder cancer, and where a causal connection with benzidine-based azo dyes had been proven or suggested, e.g. for textile dyers (Jenkins, 1978), leather dyers and shoeworkers (Decouflé, 1979), hairdressers (Guberan et al., 1985), and tailors (Anthony and Thomas, 1970). The results of our own survey of painters are very probably not relevant for the present working conditions in Germany and other highly industrialized countries, because of different materials, working methods, and hygienic standards introduced in recent years. They are, however, quite relevant for matters of compensation of persons who are now diseased. Regulatory aspects (FRG) The arguments described above have led the German Commission for Investigation of Health Hazards of Chemical Compounds in the Work Area (MAK-Commission) to include the following chapter in the MAKlist, since 1988 (DFG, 1988): Azo colorants are characterized by the azo group -N=N-. They are made by the coupling of singly and multiply diazotized aryl amines. Of particular toxicological importance are colorants from double diazotized benzidine and from benzidine derivatives (3,3′dimethylbenzidine, 3,3′-dimethoxybenzidine, 3,3′-dichlorobenzidine). In addition, aminoazo-benzene, aminonaphthalene and monocyclic aromatic amines are encountered. Reductive fission of the azo group, either by intestinal bacteria or by azo reductases of the liver and extrahepatic tissues, can cause these compounds to be released. Such breakdown products have been detected in animal experiments as well as in man (urine). Mutagenicity, which has been observed with numerous azo colorants in in-vitro test systems, and the carcinogenicity in animal experiments are attributed to the release of amines and their subsequent metabolic activation. There are now epidemiological indications that occupational exposure to benzidinebased azo colorants can increase the incidence of bladder carcinomas. Thus, all azo colorants whose metabolism can liberate a carcinogenic aryl amine are suspected of having carcinogenic potential. Due to the large number of such dyes (several hundred) it seems neither possible

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nor justifiable to substantiate this suspicion in each individual case by means of animal experimentation according to customary classification criteria. Instead, scientifically justifiable models have to be relied on. Therefore, as a preventive measure to avoid putting exposed persons at risk, it is recommended that the substances be dealt with as if they were classified in the same categories as the corresponding carcinogenic or suspected carcinogenic amines (A1, A2, B) If there are indications that the colorant itself (e.g. a pigment) or any carcinogenic breakdown products are not biologically available, the exclusion of risk should be experimentally proven or substantiated by biomonitoring. Suitable animal experiments can also rule out suspicion of carcinogenic potential. On the basis of this general view, which had been endorsed by the German Ministry of Labor (TGS 900, Bundesarbeitsblatt 1/1990, p. 63), the identification of the aromatic amine component of azo colorants is of key importance. A suitable compilation of azo colorants, according to their aromatic amine components, has been published by Myslak (1990). Azo pigments The postulate of further research on the bioavailability and/or carcinogenicity of azo pigments, especially those based on 3,3′dichlorobenzidine (v.s.), has focused interest on this particular problem. Azo pigments based on 3,3′-dichlorobenzidine (e.g. Pigment Yellow 12, Pigment Yellow 13, Pigment Yellow 14) have been orally administered to rats, hamsters, rabbits and monkeys, at doses up to 400 mg pigment kg−1 b.w. (Leuschner, 1978; Mondino et al., 1978; DHEW, 1978; Nony et al., 1980; Decad et al., 1983; Sagelsdorff et al., 1990; Hoechst AG, unpublished data). These authors could not find 3,3'-dichlorobenzidine in the urine of animals treated with 3,3'-dichlorobenzidine pigments. Decad et al. (1983) demonstrated that 14C-labelled Pigment Yellow 12, orally administered to rats, was completely excreted in the faeces. On the basis of these investigations of disposition of 3,3′-dichlorobenzidine-based pigments, it is clear why none of the long-term animal carcinogenicity studies performed so far (Leuschner, 1978; DHWE 1978; ICI, unpublished data) has demonstrated a carcinogenic effect of a diaryl pigment. It therefore appears that the aromatic amine components from azo pigments are practically not bioavailable, as demonstrated for several pigments on the basis of 3,3′-dichlorobenzidine (ETAD, 1990; see also Bolt and Golka, 1993). Hence, it is now very unlikely that occupational exposure to insoluble azo pigments would be associated with a substantial risk of (bladder) cancer in man.

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References AKIYAMA, T., 1970, The investigation on the manufacturing plant of organic pigment, Jikei Med. J., 17, 1–9. ANTHONY, H.M., 1974, Industrial exposure in patients with carcinoma of the bladder, J. Soc. Occup. Med., 24, 110–16. ANTHONY, H.M. and THOMAS, G.M., 1970, Tumors of the urinary bladder: An analysis of the occupations of 1030 patients in Leeds, England, J. Nat. Cancer Inst., 45, 879–95. BOENINGER, H., 1978, An investigation of the metabolic reduction of benzidine azo dyes to benzidine and its metabolites and their possible relationship to carcinoma of the bladder in man. Unpublished data (cited by Gregory, 1984). BOLT, H. M and GOLKA, K., 1993, Zur früheren Exposition von Malern gegenüber Azofarbstoffen, Arbeitsmed., Sozialmed., Umweltmed., 28, 417–21. Bos, R.P., VAN DER KRIEKEN, W., SMEIJSTERS, L., KOOPMAN, J.P., DEJONGE, H.R., THEUWS, J.L.G. and HENDERSON, P.T., 1986, Internal exposure of rats to benzidine derived from orally administered benzidine-based dyes after intestinal azo reduction, Toxicology, 40, 207–13. BOWMAN, M.C., NONY, C.R., BILLEDEAU, S.M., MARTIN, J.L. and THOMPSON, H.C., 1983, Metabolism of nine benzidine-congener-based azo dyes in rats based on gas chromatographic assays of the urine for potentially carcinogenic metabolites, J. Anal. Toxicol. 7, 55–60. CERNIGLIA, C.E., FREEMAN, J.P., FRANKLIN, W. and PACK, L.D., 1982, Metabolism of azo dyes derived from benzidine, 3,3'-dimethylbenzidine and 3, 3'-dimethoxybenzidine to potentially carcinogenic aromatic amines by intestinal bacteria, Carcinogenesis, 3, 1255–60. CHUNG, K.T., FULK, G.E. and EGAN, M., 1978, Reduction of azo dyes by intestinal anaerobes, Appl. Environ. Microbiol. 35, 55–62. Colour Index, 3rd edn., 3rd rev., 1987, Bradford: Society of Dyers and Colourists. Vols. 1–8. DECAD, G.M., SNYDER, C.D. and MITONA, C., 1983, Fate of water-insoluble and water-soluble dichlorobenzidine-based pigments, J. Toxicol. Environm. Hlth, 11, 455–65. DECOUFLÉ, P., 1979, Cancer risk associated with employment in the leather and leather products industry, Arch. Environm. Hlth, 34, 33–7. DFG (Deutsche Forschungsgemeinschaft), 1988, List of MAK and BAT Values 1988, Weinheim: VCH Publishers. DEWAN, A., JANI, J.P., PATEL, J.S., GANDHI, D.N., VARIYA, M.R. and GHODSARA, N.B., 1988, Benzidine and its acetylated metabolites in the urine of workers exposed to Direct Black 38, Arch. Environm. Hlth, 43, 269–72. DHEW, 1978, Bioassay of diarylanilide yellow for possible carcinogenicity, DHEW Publication No. (NIH) 78–830, US Dept of Health, Education and Welfare, Public Health Service, National Cancer Institute, Carcinogens Testing Program. ETAD (Ecological and Toxicological Association of the Dyestuffs Manufacturing Industry, 1990, Zum kanzerogenen Potential von Diaryl-Azopigmenten auf Basis von 3,3′-Dichlorbenzidin, ETAD-Bericht T 2028-BB (D), ETAD, CH-4005, Basel 5.

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FOUTS, J.R., KAMM, J.J. and BRODIE, B.B., 1957, Enzymatic reduction of prontosil and other azo dyes, J. Pharmacol, Exp. Ther., 110, 291–300. GENIN, W.A., 1977, Formation of clastogenic diphenylamine derivates as a result of the metabolism of direct azo dyes, Vopr. Oncol, 23, 50–2 (in Russian). GREGORY, A., 1984, The carcinogenic potential of benzidine-based dyes, J. Environm. Pathol. Toxicol. Oncol, 5, 243–59. GUBERAN, R., RAYMOND, L. and SWEETNAM, P.M., 1985, Increased risk for male bladder cancer among a cohort of male and female hairdressers from Geneva, Int. J. Epidemiol, 14, 549–54. HARTMANN, C.P., FULK, G.E. and ANDREWS, A.W., 1978, Azo reduction of trypan blue to a known carcinogen by a cellfree extract of a human intestinal anaerobe, Mutat. Res., 58, 125–32. JENKINS, C.L., 1978, Textile dyes are potential hazards, J. Environm. Hlth, 40, 279– 84. KENNELLY, J.C., HERTZOG, P.J. and MARTIN, C.N., 1982, The release of 4,4′diaminobiphenyls from azo dyes in the rat, Carcinogenesis, 3, 947–51. LEUSCHNER, F., 1978, Carcinogenicity studies on different diarylide yellow pigments in mice and rats, Toxicol. Lett., 2, 253–60. LOWRY, L.K., TOLOS, W.P., BOENINGER, M.F., NONY, C.R. and BOWMAN, M.C., 1980, Chemical monitoring of urine from workers potentially exposed to benzidine-derived azo dyes, Toxicol. Lett., 7, 29–36. LYNN, R.K., DONIELSON, D.W., ILIAS, A.M., KENNISH, J.M., WONG, K. and MATHEWS, H.B., 1980, Metabolism of bisazobiphenyl dyes derived from benzidine, 3,3′-dimethylbenzidine or 3,3′-dimethoxybenzidine to carcinogenic aromatic amines in the dog and rat, Toxicol Appl. Pharmacol, 56, 248–58. MARTIN, C.N. and KENNELLY, J.C., 1981, Rat liver microsomal azoreductase activity on four azo dyes derived from benzidine, 3,3′-dimethylbenzidine or 3,3′dimethoxybenzidine, Carcinogenesis, 2, 307–12. MARTIN, C.N. and KENNELLY, J.C., 1985, Metabolism, mutagenicity and DNA biding of biphenyl-based azo dyes, Drug Metab. Rev., 16, 89–117. MEAL, P.F., COCKER, J., WILSON, H.K. and GILMOUR, J.M., 1981, Search for benzidine and its metabolites in urine of workers weighing benzidine-derived dyes, Br. J. Ind. Med., 38, 191–3. MONDINO, A., ACHARI, R., DUBINI, M., MARCHISIO, M.A., SILVESTRI, S. and ZANOLO, G., 1978, Absence of dichlorobenzidine in the urine of rats, rabbits and monkeys treated with C.I. Pigment Yellow 13, Med. Lav., 69, 693– 7. MYSLAK, Z.W., 1990, Azofarbmittel auf der Basis krebserzeugender und verdächtiger aromatischer Amine. Identification, Verwendungsbereiche, Herstellungszeiträume. Schriftenreihe der Bundesanstalt für Arbeitsschutz, GA 35, Bremerhaven: Wissenschaftsverlag NW. MYSLAK, Z.W. and BOLT, H.M., 1988, Berufliche Exposition gegenüber Azofarbstoffen und Harnblasenkarzinom-Risiko, Zbl. Arbeitsmed., 10, 310– 21. MYSLAK, Z.W., BOLT, H.M. and BROCKMANN, W., 1991, Tumors of the urinary bladder in painters: a case-control study, Am. J. Ind. Med., 19, 705–13. NONY, C.R., BOWMAN, M.C., CAIRNS, T., LOWRY, L.K. and TOLOS, W.P., 1980, Metabolism studies of an azo dye and pigment in the hamster based on

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analysis of the urine for potentially carcinogenic aromatic amine metabolites. J. Anal. Toxicol, 4, 132–40. RADOMSKI, J.L. and MELLINGER, T.J., 1962, The absorption, fate and excretion in rats of the water-soluble azo dyes, FD&C Red No. 2. FD&C Red No. 4 and FD&C Yellow No. 6, J. Pharmacol. Exp. Ther., 136, 259–66. RINDE, E. and TROLL, W., 1975, Metabolic reduction of benzidine azo dyes to benzidine in the Rhesus monkey, J. Nat. Cancer Inst., 55, 181–2. ROBENS, J.F., DILL, G.S., WARD, J.M., JOINER, J.R., GRIESEMER, R.A. and DOUGLAS, J.F., 1980, Thirteen-week subchronic toxicity studies of Direct Blue 6, Direct Black 38 and Direct Brown 95 dyes, Toxicol. Appl. Pharmacol., 54, 431–42. SAGELSDORFF, P., JOPPICH-KUHN, M. and JOPPICH, M., 1990, Biomonitoring for the bioavailability of dichlorobenzidine from dichlorobenzidine-based dyes, J. Cancer Res. Clin. Oncol, 116, 79 (abstract). WALKER, R., 1970, The metabolism of azo compounds, a review of the literature, Food Cosmet. Toxicol, 8, 659–76. YOSHIDA, O., HARADA, T., MIYAKAWA, M. and KATO, T., 1971, Bladder cancer among dyers in the Kyoto area, Igaku Ayumi, 79, 421–2 (in Japanese). YOSHIDA, O. and MIYAKAWA, M., 1973, Etiology of bladder cancer: metabolic aspects, in Nakahara, W., Hirayama, T., Nishioka, K. and Sugano, H. (Eds) Analytical and Experimental Epidemiology of Cancer, Proc. of the 3rd Int. Symp. of Princess Takamatsu Cancer Research Fund, pp. 31–9, Baltimore: University Park Press.

22 Toxicology of Textile Chemicals DIETER SEDLAK EnviroTex GmbH, Augsburg

The former main aspects of textiles like fashion or usefulness seem to be pushed back by a new phenomenon, the ecological and toxicological aspect of textiles. Although everybody is talking about textiles in this context, everybody means the textile chemicals (and dye-stuffs) on the textile. With respect to this situation we have to consider new developments in Germany (Figure 22.1). For two years now textile finishing plants have to be approved within the German federal immission control act. This means that all sorts of immissions to the working place or the surroundings of the plant must be defined in quantity and quality. In the meantime so-called emission factors have been developed. But this project is not yet finished. Even for insiders it was surprising how many textile chemicals show unexpected behaviour during their application. The reasons are: – – – –

unknown byproducts or impurities, unknown reactivities between components, unknown interactions with substrate, unknown dependencies of process parameters.

Today producers and users realise that the inherent toxicological properties of many textile chemicals evaluated within a typical safety data sheet represent only a part of the whole knowledge you need for safe handling and processing. A further interesting development undoubtedly is the discussion around the ecolabelling of textiles. There definitely is some danger so that different labels based on different commercial interests should be evaluated. A positive step could be the fact that MST and Ecotex have joined. But other societies like TÜV or GSF create their own labels including statements that they use the better (right) label criteria. This leads quite clearly to total confusion for the consumer. However, what the textile industry needs today is confidence. The developments discussed above will not help. From a technical point of view we have the same problem as discussed with respect to the immissions. Too little is known about the real properties and

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Figure 22.1 New developments in Germany regarding toxicology of textile chemicals.

behaviour of textile chemicals to define the absolute label criteria. These uncertainties can be solved only by cooperation of the different groups and not by aggressive competition. How can we handle both developments in a proper way? At first we have to be clear about the interactions of both problem fields. Let us start with the textile chemical delivered to a finishing plant (Figures 22.2 and 22.3). Following the directives for dangerous substances or safety data sheets the product is exactly labelled and its toxicological

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Figure 22.2 Possible levels to discuss the toxicology of textile chemicals.

properties are well described in the SDS. This is only as good as the information given to the product safety manager responsible for the possible ways of handling and processing. However, in most cases textile chemicals can be described as harmless. Nearly all show an acute oral toxicity greater than 2000 mg kg−1 although they may contain toxic substances in diluted form. Only a few of them possess an irritant character or even CMT properties. Nevertheless, many of them have impurities with these properties which are given more and more, even in low concentrations. This is a very important process

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Figure 22.3 Toxicological profile of textile chemicals.

because these low concentrations may also lead to high vapour concentrations at the workplace if these substances are volatile. This is mostly the case. For example, the carcinogenic substance acrylonitrile has a product label only giving a concentration of 0.2 percent and more in the corresponding polymer dispersion. Even 0.02 percent may lead to workplace concentrations a hundred fold higher than the TLV allows. Product data sheets with these ‘properties’ still exist with no additional information because it is not needed! Another problem is that nearly all textile chemicals are exposed to temperatures between 100°C and 230°C during application which leads to many only poorly defined substances which may all be set free in the workplace and the surroundings. Here we are also confronted with a typical juristic phenomenon. Many suppliers do not take any responsibility regarding these questions if the user defines his process parameters himself, especially when using recipe components from different suppliers. But everybody knows that the user has not the means to evaluate the toxicological profile of his chemicals mixture and process. This lack of responsibility should be clarified. Now let us observe the result of such chemical treatment of textiles. Normally you will only discuss the summarised toxicological profiles of the used chemicals in an additive way. This means that the combination of different products—all with no acute toxicity—will also show no acute

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Figure 22.4 Typical composition of a flame retardant recipe.

toxicity. Furthermore the concentration of the active ingredients in the new matrix textile is much lower than in the matrix water most textile chemicals are based on. This strategy seems to be appropriate. On the other hand take the irritant property of, for example, fatty amine based emulsifier. This property is lost during formulation of the emulsifier in the textile chemical by homogeneous dilution. The average concentration after application to the textile is surely lower. But who has information about the actual form of this fatty amine in the textile. Is it distributed homogeneously, is it more located at the surface, does it interact with other substances and even lose its irritant character? Many questions seem to be unanswered. Both complex questions—the toxicology of the handling and processing of textile chemicals and the toxicology of the result of the processing on textiles— will be discussed by means of an admittedly drastic example, a flame retardant process. This example is rather representative because an

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Figure 22.5 Total composition of a flame retardant recipe.

EEC-directive is ‘under construction’ just now which demands the finishing of all upholstery covers with flame retardant chemicals. The composition of this recipe, shown in Figure 22.4, doesn’t seem to be too complex. No special toxicological properties are apparent. The corrosive character of the phosporic acid vanishes during the application

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Figure 22.6 Release of chemical substances during/after a flame retardant process.

by neutralizing the finished textile with alkalis. Complex reactions of the different components are expected which are supposed to form not well defined polycondensates fixed to the textile fibre. On this basis there is no apparent reason to think of any toxicological side effects. In Figure 22.5, however, you will get a good idea about the real situation. But this detailed composition should not be deceptive about the fact that this is the ‘wanted’ technical composition verified by a few analytical data. In this case about 500 actual substances can be expected. Even if we should have available all the necessary toxicological data for the components it would not help, because there is poor information about the result of the finishing process itself. This is the only reality. A first step in the right direction would be the analysis of the process offgas. The result reflects approximately the activities on the textile during

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Figure 22.7 Needs for the textile toxicology assessment.

processing. Figure 22.6 shows in which way this knowledge can be used to perform a proper risk assessment based on defined substances. However, in the case of evaluation of the skin toxicity we are confronted with a lack of data concerning the bio-availability of the named substances. Not before having cleared these additional questions can the toxicologist seriously start his routine activities, the risk assessment. Anyway this seems to be much easier than to discover the chemical basis for this assessment. With respect to the toxicology of textile chemicals there is a lot to do mostly in the field of understanding the chemicals and processes used. After this we recommend the development of models for evaluating the bioavailability of substances and test kits to characterize the toxicological properties of textiles as a whole. Last but not least we hope for much better communication between the industries involved (Figure 22.7).

23 Antioxidants and Light Stabilisers : Toxic Effects of 3,5-Di-alkyl-4hydroxyphenyl Propionic Acid Derivatives in the Rat and their Relevance for Human Safety Evaluation HELMUT THOMAS, PETER DOLLENMEIER, ELKE PERSOHN, HANSJÖRG WEIDELI and FELIX WAECHTER Ciba-Geigy Limited, Basle Introduction Antioxidants and light stabilisers are important additives for a wide range of plastic materials used for industrial as well as for medicinal and food packaging purposes. Technical efficiency requires that these compounds are mobile within the polymer network. This implies that humans may be exposed to such compounds not only during the manufacturing process but also in the course of the decay of polymers, by direct contact with substances that have migrated to the surface of plastic materials or by ingestion of diffusion-contaminated food. It is of considerable interest, therefore, to be aware of the toxicological properties of these compounds and the relevance of these properties for the safety assessment in humans. Sterically hindered phenolic antioxidants Phenolic antioxidants have been widely used as food preservatives and their almost classical representatives, 2,6-di-tert-butyl-4-methyl phenol (BHT) and 2-tert-butyl-4-methoxyphenol (BHA) have been characterised in extenso with respect to their biochemical and toxicological properties (Søndergaard and Olsen, 1982; Conning and Phillips, 1986; Ito et al., 1986a; Williams et al., 1990a,b; Verhagen et al., 1991). Being nongenotoxic, these compounds were found upon oral administration to rodents to cause slightly increased liver weight, to induce mainly epoxide hydrolase and phase II drug metabolising enzymes and to exert some anticarcinogenic effect (Benson et al., 1979; Choe et al., 1984; Gregus and Klaassen, 1988; Prochaska and Talalay, 1988; Perchellet and Perchellet, 1989; Rodrigues et al., 1991). Pulmonary toxicity and carcinogenicity particularly of BHT in the mouse and formation of forestomach carcinoma and papilloma in the rat and Syrian hamster, respectively, have been reported and are considered to be largely species-specific effects (Abraham et al., 1986; Anon, 1986; Ito et al., 1986a, b; Verhagen et al., 1989). Industrial phenolic antioxidants as used in the polymer technology, by

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contrast, although related to BHT, require higher molecular weights and hydrophobicity in order to retain them in the polymer matrix. This can be achieved, for example, by introducing alkyl chains or (esterified) carboxyalkyl moieties in the para-position to the phenolic hydroxyl group with variations in the aliphatic substitution pattern of the 2- and 6positions. With these modifications the questions arose, whether or not the toxicology of the chemically modified offspring would still be related to that of the phenolic core of the BHT ancestor or completely different, although a common metabolic degradation pathway to structural entities resembling BHT could be anticipated. Clarification of this question was expected to contribute to the understanding of the toxicology of an entire class of phenolic antioxidants. A number of differentially esterified 4hydroxy-3, 5-dialkyl-phenylpropionic acid derivatives were subsequently subjected to subchronic toxicity testing in rats with the result that the effects encountered were largely dependent on the alcohol moiety used for esterification and the size of the alkyl-substituent in the 3- and 5-positions: most frequently, increased liver weights were encountered in compounds with bulky 3,5-substituents such as tert-butyl. In some instances, however, initial hepatomegaly was followed after several weeks or months of treatment by increases in thyroid weights and a proliferation of the thyroid follicular epithelium. The mechanism leading to the latter effect has been investigated in detail using the di-ester of 3-tert-buty1–4-hydroxy-5-methylphenylpropionic acid with ethylene glycol, Compound B, as a model compound (Table 23.1). Blood kinetics and blood metabolites Compound B is a symmetrical di-ester compound (Table 23.1). When administered at a single oral dose of about 10 mg kg−1 body weight to male rats, 14C-phenyl-labelled Compound B was readily adsorbed, and maximal blood radioactivities were reached after 1 h. Thereafter, blood radioactivity declined rapidly and only minute amounts were detected 48 h after treatment. At any time point investigated Compound A, the free carboxylic acid of Compound B, was the dominating blood metabolite, whereas the parent compound constituted a minor component only (Table 23.2). These findings are indicative of an efficient first pass hydrolysis and suggest that the carboxylic acid metabolite, Compound A, might be responsible for the toxicological profile of this antioxidant in the rat. Liver enzyme induction Compound B was administered in the feed to male rats at dose levels of 50, 150, 500 and 1000 ppm. After treatment for 14 days, absolute liver weights were dose-dependently increased. Biochemically, this

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Table 23.1 3.5-Substituted 4-hydroxyphenyl propionic acid esters

hepatomegaly was accompanied by an induction of total microsomal cytochrome P450, microsomal epoxide hydrolase and UDPglucuronosyltransferase, cytosolic glutathione S-transferase and peroxisomal β-oxidation. The observed induction of the cytochrome P450 content was reflected in increased activities of ethoxycoumarin O-deethylase as well as lauric acid 11- and 12-hydroxylases (Table 23.3). This suggests an induction of cytochrome P450 isoenzymes of the subfamilies CYP2B and CYP4A. Indeed, increased amounts of CYP2B and CYP4A proteins were found in liver microsomes from treated animals by means of Western-blotting with monoclonal antibodies specific for these iso-enzymes (data not shown). Within 28 days after cessation of a 14-day treatment at

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Table 23.2 Parent equivalents and blood metabolites of [14C]-labelled Compound B after single oral administration of 9.5 mg kg−1 body weight to male rats

Note: bld: below the limit of detection. Blood was taken at the indicated time intervals and extracted with ethyl acetate for analysis. Compound B and its free carboxylic acid metabolite (Compound A, Table 23.1) were identified by thin-layer co-chromatography with authentic reference samples. Quantification of Compounds A and B was accomplished by radiometric scanning of the plates following thin-layer chromatography of the respective blood extracts.

1000 ppm, liver weights as well as the investigated biochemical parameters returned to control levels. Therefore, Compound B may be addressed as a reversible barbiturate- and peroxisome proliferator-type inducer in the rat, as characterised by its liver enzyme induction profile. Effects on serum thyrotropin and thyroid hormones When male rats were fed Compound B admixed in the diet for 14 days at dose levels of 50, 150, 500 and 1000 ppm, liver and thyroid weights were increased in a dose-dependent manner. Histopathological examination of the thyroid gland revealed hypertrophy of the follicular epithelium and thinning of colloid at 150, 500 and 1000 ppm. Morphological changes in the pituitary gland comprised enlarged thyrotropin (TSH)-producing cells with foamy or vacuolated cytoplasm. In addition, treatment resulted in markedly increased serum TSH and reverse triiodothyronine (rT3) concentrations, whereas serum thyroxine (T4) and triiodothyronine (T3) levels were found slightly decreased. The effects observed at 1000 ppm were found to be reversible after a 28-day recovery period (MuakkassahKelly et al., 1991). In additional experiments, male rats were rendered hypothyroid, fed Compound B at 1000 ppm for 21 days and infused for the last 7 days with slightly supraphysiological concentrations of T4. The observed changes in

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Table 23.3 The effect of Compound B on selected biochemical parameters in the male rat liver

Notes: Values are means of 9 (14 days treatment) or 4 (14 days treatment/28-days recovery) animals. Standard deviations are given in parentheses. 0/0: Control recovery. 1000/0: High-dose recovery. * p<0.01.

blood thyroid hormone levels were typical of an inhibitor of deiodinase type I (5′D-I) activity: slightly increased T4 levels were accompanied by markedly decreased T3 levels and markedly increased rT3 levels. 5′D-I activities and mRNA levels were measured in the liver and kidney of these animals and found markedly inhibited and decreased, respectively (Liang et al., 1993). These results suggest that Compound B, by inhibiting 5′D-I activity, decreases T3 levels and, since the expression of 5′D-I activity is regulated by T3, secondarily also 5′D-I mRNA levels. Compound B obviously interferes with thyroid hormone homeostasis in rats. Treated animals compensate for this disturbance with elevated TSH levels, which in turn induce thyroid follicular hypertrophy, in order to maintain an euthyroid state. The available experimental results suggest that Compound B causes this effect indirectly by altering peripheral, in particular hepatic thyroid hormone metabolism: the compound inhibits 5′deiodination of T4 to the more potent hormone T3. In addition, the observed induction of liver UDP-glucuronosyltransferase (Table 23.3) is indicative of an increased formation of T4-glucuronide, the rate-limiting step for biliary excretion of this hormone (Thomas and Williams, 1991). Thyroid hypertrophy as a consequence of altered thyroid hormone

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metabolism has been observed with various deiodinase inhibitors (Hill et al., Liang et al., 1993) and hepatic enzyme inducers (McClain, 1989; Curran and DeGroot, 1991; Barter and Klaassen, 1992; Visser et al., 1993). Induction of thyroid neoplasia In a long-term feeding study, administration of Compound B to rats at a dose level of 1000 ppm was associated with an increased incidence of thyroid gland follicular adenoma and carcinoma (Muakkassah-Kelly et al., 1991). Compound B was shown to be devoid of mutagenic and clastogenic activity (Muakkassah-Kelly et al., 1991). Therefore, it is likely that thyroid tumour induction by Compound B was not the result of a direct, genotoxic effect on this organ, but rather a consequence of the hormonal imbalance induced by this antioxidant in the rat. An intact hypothalamic-pituitary-thyroid axis is able to respond to a chemically induced alteration in peripheral hormone metabolism with increased hormone production by the hypertrophic gland. However, it is known that chronic, excessive stimulation of the gland can lead to follicular hyperplasia and ultimately progress to thyroid neoplasia (Paynter et al., 1988; McClain, 1989; Curran and DeGroot, 1991; Johnson et al., 1993). For Compound B, this hypothesis is in agreement with the doseresponse characteristics obtained in the long-term study, where thyroid tumours were induced exclusively at a dose-level sufficiently high to cause hormonal imbalance (Muakkassah-Kelly et al., 1991). Implications for human risk assessment For human risk assessment, it is of critical importance to identify the mechanism by which Compound B caused thyroid neoplasia in the rat, a species most often used for carcinogenic hazard identification. Hyperplastic changes in the thyroid are frequently observed in rat carcinogenicity studies, and this species appears to be very sensitive to compounds which interfere with thyroid hormone synthesis and/or catabolism. They evoke an immediate stimulation of the gland upon short-term treatment as a consequence of an increased pituitary TSH secretion (Zbinden, 1987; Paynter et al., 1988; McClain, 1989). However, the effects of xenobiotics on the pituitary-thyroid axis in rodents cannot necessarily be extrapolated to man since rodents and man are distinguished by many important physiological and biochemical differences (Gopinath, 1991): e.g. the amount of thyroxin-binding globulin, the half-life of T4 and its biliary excretion as well as plasma THSlevels and their response to thyrotropin releasing hormone. These differences render the rat very sensitive to small changes in the plasma T4

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level, whilst humans are essentially insensitive (Hill et al., 1989; Grasso et al., 1991). Therefore, in contrast to rats, there is no conclusive evidence for a critical role of TSH in thyroid stimulation and carcinogenesis in humans (Hill et al., 1989). In cultured human thyroid cells, for example, THS was unable to induce proliferation whereas a stimulation of growth was observed in rat thyroid cells (Westermark et al., 1985). Clinical data are available for some compounds such as the anticonvulsants phenobarbitone, diphenylhydantoin and carbamazepine as well as the antibiotic rifampicin, which are known liver microsomal enzyme inducers in man. They increase thyroid hormone metabolism and excretion and eventually decrease serum thyroid hormone levels. There is also evidence, that administration of these drugs leads to thyroid stimulation, however, largely in the absence of increased TSH levels (Curran and DeGroot, 1991). In addition, epidemiological data are not in favour of a link between human use of such compounds with an increased incidence of thyroid tumours (Curran and DeGroot, 1991), nor have increased rates of thyroid cancers been reported in areas of endemic iodine deficiency (McClain, 1989). Therefore, the currently available data do not support the idea, that thyroid stimulation as a response to chemically induced increases in circulating TSH concentrations significantly contributes to thyroid tumour formation in man. Compound B has been shown to cause thyroid tumours in the rat. In a series of speciality studies, the compound was identified as an enzyme inducer and a 5′-deiodinase inhibitor in the rat liver. These findings argue for a rodentspecific, indirect mechanism leading to the formation of thyroid tumours. Moreover, the observed dose-response characteristics are indicative of a threshold process, e.g. liver enzyme induction and inhibition of 5′-deiodination are irrevocable prerequisities for thyroid tumour formation in this species. Benzotriazole-based light stabilisers Ester derivatives of the 3-[3-(2H-benzotriazole-2-yl)-5-tert-butyl-4hydroxy-phenyl] propionic acid represent potent UV-light absorbers and constitute an important class of industrial plastic additives and light stabilisers (Compounds C-F, Table 23.1). Toxicologically, this class of chemicals is characterised by generally low acute oral or dermal toxicity and the lack of genotoxicity in the commonly employed battery of bacterial and cellular mutagenicity tests. Irrespective of the alcohol moiety, however, all compounds, when administered subchronically to rats, displayed very similar predominantly hepatotrophic effects, with spleen and kidney weights in addition being only slightly affected: pronounced hepatomegaly, hepatocyte hypertrophy, and concomitantly increased

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plasma transaminase activities. Upon electron microscopical examination, a striking peroxisome proliferation was the major finding. Compound F, a di-ester ‘product by process’ obtained upon esterification of 3-[3-(2H-benzotriazole-2-yl)-5-tert-buty1–4-hydroxyphenyl] propionic acid with polyethyleneglycol 300, when tested for its toxicity to rat reproduction in a Segment I study, gave rise to increased numbers of stillborn pups, decreased pup survival, decreased weight gain of surviving pups, and dark discoloured abdominal skin regions in a number of pups at higher dose levels. The common nature of general toxicology findings with all investigated derivatives of the addressed benzotriazole-based light stabilisers suggested a common basis of action and, depending upon this action, perhaps a very similar behaviour and extent of potency as foetotoxic agents. In order to investigate these interrelationships a series of mechanistic studies were conducted focusing on the kinetics, primary metabolism of the parent compounds in vitro and in vivo and their effect on selected biochemical liver parameters in rats. Compound F was selected as a model compound to investigate the mechanism of toxicity in pregnant female rats and foetuses. In vitro hydrolysis Compound D, the methyl ester of 3-[3-(2H-benzotriazole-2-yl)-5-tertbutyl-4-hydroxyphenyl] propionic acid was readily hydrolysed in vitro by rat serum as well as rat liver homogenate while a homogenate of rat small intestine when compared on a gram tissue basis, appeared to be less efficient by three orders of magnitude than the liver. Increasing the sterical hindrance around the ester bond by formation of the di-ester with a short chain alcohol reduced the rate of in vitro hydrolysis considerably as demonstrated for Compound F, the diester of hexane-l,6-diol with Compound C. When offered at a test concentration of 0.2 mM, essentially no hydrolysis of this compound was observed with rat serum, and the hydrolysis by liver and small intestine homogenate was estimated to proceed at least two and one orders of magnitude slower, respectively, than calculated for Compound D (Table 23.1 and 23.4). Blood kinetics and blood metabolites Assuming comparable extents of intestinal absorption in vivo, the observed differences in the in vitro hydrolysis rates might as well suggest significantly different in vivo hydrolysis rates and consequently quite different residence times for both parent compounds in the rat in vivo. Different residence times, on the other hand, may eventually allow not only for additional routes of metabolism but also for an intensification of toxic

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Table 23.4 Kinetic parameters for the in vitro hydrolysis of Compound D and E by rat serum and organ homogenates

Notes: a The apparent V −1 ml−1. max value for serum is given in µmol min b Initial velocity of hydrolysis at 0.2 mM ester concentration. Hydrolysis of Compound D was determined in 50 mM Tris/phosphate buffer, pH 7. 5, containing either 1 per cent (v/v) rat serum, or 1.25 per cent (w/v) rat liver homogenate or 10 per cent (w/v) small intestine homogenate. Similarly, hydrolysis of Compound E was assessed using 98 per cent (v/v) rat serum or in the presence of 10 mM Tris/HCl buffer, pH 7.5, containing 250 mM sucrose and either 24.5 per cent (w/v) rat liver homogenate or 19.6 per cent (w/v) small intestine homogenate.

effects. This question was addressed in a pharmacokinetic study under conditions of single oral administration of Compounds D and E at a dose level of 10 mg kg−1 each (Table 23.5). 14C-Phenyl-labelled Compound D was readily absorbed from the gastrointestinal tract. Maximal blood radioactivity was reached between 1 and 2 h and subsequently eliminated with an apparent half-life of 10.0–11. 8 h. After 48 h only minute amounts of radioactivity equalling about 3 per cent of the blood levels in Tmax were detectable. Analysis of the resulting blood metabolite pattern largely confirmed the findings of the preceding in vitro investigation of enzymatic Compound D hydrolysis: particularly during periods of high parent equivalent concentrations in blood as recorded between 30 min and 6 h after dosing, hydrolysis appeared to be the major metabolic pathway as evidenced by the high concentrations of the carboxylic acid, Compound C (34–77 per cent), and an unidentified metabolite (17–36 per cent) which was regarded to have evolved from Compound C by an additional metabolic step (Table 23.5). Quite surprisingly, Compound E was found to be absorbed to a much lower extent than Compound D, with a Cmax after 1 h of less than one tenth of the value seen with the latter. Elimination with an apparent halflife of 12.0 h and slightly higher residual radioactivity of approximately 4.8 per cent of the blood levels recorded at Tmax after 48 h indicated only a slightly reduced elimination rate as compared to Compound D. Also, quite different from the initially anticipated result, hydrolysis contributed substantially to the rapid metabolism of Compound E. The 24 h AUC values revealed a 23 per cent and 36 per cent contribution of the carboxylic

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Table 23.5 Summary of pharmacokinetic parameters obtained after single oral administration of 10 mg kg−1 Compound D and E to two male rats each

Notes: a Below the limit of reliable quantification. pe: Parent equivalent.

acid, Compound C, and unidentified metabolites, respectively, to the total AUC. It is assumed that the majority of unidentified metabolites have arisen from further biotransformation of the carboxylic acid. The significant contribution of hydrolysis products to the total AUC is still evident after 168 h although low blood radioactivity levels 24 h after dosing generally prevented accurate quantitation of the metabolites and thus slightly diminished their apparent overall share (Table 23.5). Liver enzyme induction Subchronic oral (gavage) administration of single daily doses of Compound C for 14 days, Compound D for 14 days, Compound E for 13 weeks, and Compound F for 114 days to male rats (Tables 23.6 and 23.7) was correlated with a dose-dependent massive increase in absolute liver weight up to about 190 per cent of control at the highest dose level irrespective of the treatment period and the test compound. This pronounced hepatomegaly was paralleled by a comparably small two-fold elevation of the microsomal cytochrome P450 contents and an about 50 per cent decrease in total UDP- glucuronosyltransferase activity. Essentially no changes were recorded for the cytochrome P450 dependent ethoxycoumarin O-de-ethylase activity while microsomal epoxide hydrolase activities appeared to vary, with slight increases in Compound C and Compound D treated animals, no changes in Compound E treated animals and even a dose-dependent reduction to 46 per cent of control in rats treated with Compound F at 100 mg kg−1 (Table 23.6). Strongly induced peroxisomal fatty acid β-oxidation activities for all model compounds as well as lauric acid 12-hydroxylase activities tested for Compounds E and F were accompanied by significant dose-dependent decreases in glutathione S-transferase activities to 32, 51, 40 and 15 per cent of control at the highest dose level tested for Compound C, D, E and

Notes: dnt: dose level not tested, nd: not determined. Microsomal epoxide hydrolase and total microsomal UDP-glucuronosyltransferase activities were determined with styrene oxide and 3-methyl-2-nitrophenol as substrate, respectively. Values are means±standard deviation of 10 control and 5 treated animals (a, b) or 8 animals (c) or 5 animals per group (d). Asterisks indicate results significantly different (two-sided Dunnett’s test) from control: * p<0.05, † P<0.01, p<0.001.

Table 23.6 The effect of various benzotriazole-based light stabilisers on selected biochemical parameters related to and indicative for a barbiturate and/or polycyclic aromatic hydrocarbon type enzyme induction in male rat liver

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F, respectively. For Compound F, which appeared to be the most potent inducer of peroxisomal β-oxidation and lauric acid 12-hydroxylase activities, a concomitant strong 90 per cent reduction of morphine UDPglucuronosyltrasferase and marked 2.5-fold increase in bilirubin UDPglucuronosyltransferase activity was recorded (Table 23.7). Electron microscopy confirmed what had already been indicated by changes in the investigated enzyme levels, a striking proliferation of peroxisomes with the same appearance of these organelles regardless of the compound tested: vigorous increase in number, a striking number of markedly enlarged peroxisomes frequently containing matrical inclusions (matrical plates) and peroxisomes forming arrays or clusters (polyperoxisomes) in the virtual absence of any significant proliferation of smooth endoplasmic reticulum (data not shown). Consequently, the hepatotrophic effects of the tested benzotriazole-based light stabilisers were clearly assigned to their action as peroxisome proliferators in rat liver. Mindful of the different durations of treatment their potency was found to rank in the order Compound E
Notes: dnt: dose level not tested. nd: not determined. Cyanide-insensitive peroxisomal fatty acid -oxidation and cytosolic glutathione S-transferase activities were determined with [l-14C]-palmitoyl-CoA and1-chloro2,4-dinitrobenzene as substrate, respectively. Values are means±standard deviation of 10 control and 5 treated animals (a, b) or 8 animals (c) or 5 animals per group (d). Asterisks indicate results significantly different (two-sided Dunnett’s test) from control: * p<0.05, † P<0.01, ‡ p<0.001.

Table 23.7 The effect of various benzotriazole-based light stabilizers on selected biochemical parameters related to and indicative for a peroxisome proliferator type enzyme induction in male rat liver

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Table 23.8 Morphological changes of dam and foetal hepatocyte organelles after treatment with Compound F from day 6 through days 14, 17 and 20 of gestation

33 0 ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT

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biochemical investigations which revealed a moderate four-fold induction of peroxisomal β-oxidation in dams and an up to 15-fold induction of this activity at day 21 of gestation in foetal livers, respectively, with the final foetal activity exceeding dam activity by 40 per cent. Strong increases in lauric acid 11- and 12-hydroxylation, which are known to be associated with isoenzymes of the cytochrome P450 CYP4A gene family and the phenomenon of peroxisome proliferation in rodents, of up to 2.6- and 10. 5-fold in dams and 11.5- and 23.2-fold in foetuses, respectively, at day 21 of gestation were recorded as well, indicating again a slightly higher final activity in foetal than in dam liver. By contrast, catalase, known as a detoxifying enzyme for hydrogen peroxide generated particularly in the course of increased peroxisomal activity, was shown to be induced up to 5. 6-fold in dam and 8.2-fold in foetal liver at day 21 of gestation, leaving foetal liver, however, with only about 30 per cent of the final activity seen in dams (Table 23.9). This discrepancy between the strong induction of hydrogen peroxide generating peroxisomal β-oxidation and the considerably less potent induction of the hydrogen peroxide destroying catalase activity appears to be reflected in the 1.7-fold increase in the lipid peroxidation product malondialdehyde in foetal livers as well as in the decrease of total and reduced hepatic glutathione levels at day 21 of gestation (Tables 23.9 and 23.10). Surprisingly, foetal defense systems against oxidative stress such as selenium-dependent and seleniumindependent glutathione peroxidase were found poorly developed throughout the investigated periods of gestation and barely inducible by the test article leaving the pups with little protection against any kind of oxidative insult (Table 23.10). Consequently, Compound F was clearly identified as a peroxisome proliferator in pregnant rat as well as foetal liver with high potential for the initiation of oxidative damage in foetal tissues. Implications for human safety assessment The understanding of the mechanisms by which benzotriazole-based UV light stabilisers exert their hepatotrophic effects in rodents is of crucial importance for the assessment of safety aspects in humans. The presented rat studies have shown that the liver effects exerted by Compound C and its ester derivatives Compounds D, E and F are clearly related to the induction of peroxisome proliferation. The identical nature of the observed effects regardless of the alcohol component in the investigated esters suggests that the toxic potential resides solely with the 3-[3-(2Hbenzotriazole-2-yl)-5-tert-butyl-4-hydroxy-phenyl] propionic acid whereby the individual potency appears to be mainly determined by the different bioavailablity of the respective ester compound and the extent and velocity of its hydrolysis in vivo. Also, it appears, that in vitro hydrolysis studies do

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not actually reflect the hydrolytic capacity of the in vivo system for a given ester. Liver enlargement and the induction of diagnostic enzyme activities is a characteristic response of rodents to treatment with peroxisome proliferators and results from a combination of both hypertrophy and hyperplasia. According to current opinion, peroxisome proliferation and liver growth are closely associated with the formation of hepatocellular tumours in rats and mice (Hawkins et al., 1987; Lock et al., 1989; Bentley et al., 1993). However, a number of feeding studies have demonstrated that there may actually be two types of threshold with respect to dose relationships: at very low doses, administration of peroxisome proliferators will not result in any liver response at all (Bentley et al., 1993). With increasing doses the first threshold will be exceeded with subsequent stimulation of peroxisome proliferation and DNA synthesis. As a result of several studies it is obvious that a limited extent of liver growth does not automatically lead to tumour formation as has been demonstrated, for example, with fenofibrate and diethylhexylphthalate in carcinogenicity assays (Mitchell et al., 1985; Price et al., 1986; Keith et al., 1991). Thus, a second threshold has to be exceeded at which the magnitude of effects is sufficient to cause tumour development in rodents. In addition, extended administration of the peroxisome proliferator appears a necessary prerequisite to exceed this tumourigenic threshold. Also, a large number of in vitro and in vivo studies have provided ample evidence for a marked species difference in susceptibility to the effects of peroxisome proliferators. Rats and mice are extremely sensitive while hamsters show a markedly smaller response and non-human primates and humans appear to be insensitive or non-responsive (Lake et al., 1989; Bentley et al., 1993; Graham et al., 1994). The latter finding is supported by epidemiological evidence from long-term treatment of patients with hypolipidaemic agents (Bentley et al., 1993). Therefore the available evidence strongly supports the conclusion that the effects of benzotriazolebased light stabilisers in rodents are of no relevance to human safety assessment. The action of Compound F as a strong peroxisome proliferator in foetal livers starting as early as day 15 of gestation suggests that treatment related initiation of high level oxidative stress under conditions of poorly developed foetal protection and detoxification systems. This view is supported by substantially elevated hepatic malondialdehyde levels and essentially depleted glycogen stores in hepatocytes of foetuses from treated dams on day 21 of gestation. The latter indicates an extensive glucose consumption, presumably via the pentose phosphate pathway, to supply the hydrogen peroxide and lipid peroxide detoxifying glutathione peroxidase system with the necessary reduction equivalents (NADPH). Under conditions of limited degradation of hydrogen peroxide, this

Notes: bld: below the limit of detection. Cyanide-insensitive peroxisomal fatty acid -oxidation and catalase activities were determined with [lC]-palmitoyl-CoA and hydrogen peroxide as substrate, respectively. Values are means±standard deviation from 6 dams per group and 6 pools of 2–9 foetuses per dam depending on the litter size. Asterisks indicate results significantly different (two-sided Dunnett’s test) from control: * p<0.05, † P<0.01, ‡ p<0.001.

Table 23.9 The effect of Compound F on dam and foetal absolute liver weight and selected biochemical liver parameters related to and indicative for peroxisome proliferation after treatment from day 6 through days 14, 17 and 20 of gestation

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Notes: Selenium-dependent and selenium-independent glutathione peroxidase activities were determined with cumene hydroperoxide and hydrogen per-oxide as substrate, respectively. Values are means±standard deviation from 6 dams per group and 6 pools of 2–9 foetuses per dam depending on the litter size. Asterisks indicate results significantly different (two-sided Dunnett’s test) from control: * p<0.05, † p<0.01, ‡ P<0.001.

Table 23.10 The effect of Compound F on selected biochemical dam and foetal liver parameters related to defence mechanisms against oxidative stress after treatment from day 6 through days 14, 17 and 20 of gestation

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compound is known to pass the liver and to be systemically distributed throughout the foetal body. Thus the endothelial cell injuries observed in the course of the Segment II study are regarded secondary to the effect of peroxisome proliferation and to have arisen as a consequence of oxidative damage which has been demonstrated to occur in various endothelial cell systems, for example as a result of iron-mediated oxygen free radical attack (Brieland et al., 1992; Barchowsky et al., 1994; Krautschick et al., 1995). Therefore, the observed foetotoxicity of Compound F in rats is regarded as occurring solely as a consequence of and secondary to peroxisome proliferation, and there is no evidence to assume that this effect as a result of exposure to benzotriazole based light stabilisers may occur in species that are non-responsive to the action of peroxisome proliferators as stated above, including humans. References ABRAHAM, R., BENITZ, K.F., PATIL, G. and LYON, R., 1986, Rapid induction of forestomach tumors in partially hepatectomized Wistar rats given butylated hydroxyanisole, Experimental and Molecular Pathology, 44, 14–20. ANON., 1986, Butylated hydroxyanisole (BHA), IARC Monographs on the Evaluation of the Carcinogenic Risk of Chemicals to Man, Vol. 40, pp. 123– 59. BARCHOWSKY, A., WILLIAMS, M.E., BINZ, C.C. and CHEPENIK, K.P., 1994, Oxidant-sensitive protein phosphorylation in endothelial cells, Free Radical Biology and Medicine, 16, 771–7. BARTER, R.A. and KLAASSEN, C.D., 1992, Rat liver microsomal UDPglucuronosyltransferase activity toward thyroxin: characterisation, induction and form specificity, Toxicology and Applied Pharmacology, 115, 261–7. BENSON, A.M., CHA, Y.-H., BUEDING, E., HEINE, H.S. and TALALAY, P., 1979, Elevation of extrahepatic glutathione S-transferase and epoxide hydratase activities by 2(3)-tert-butyl-4-hydroxyanisole, Cancer Research, 39, 2971–7. BENTLEY, P., CALDER, I., ELCOMBE, C., GRASSO, P., STRINGER, D. and WIEGAND, H.-J., 1993, Hepatic peroxisome proliferation in rodents and its significance for humans, Food and Chemical Toxicology, 31, 857–907. BRIELAND, J.K., CLARKE, S.J., KARMIOL, S., PHAN, S.H. and FANTONE, J. C., 1992, Transferrin: a potential source of iron for oxygen free radicalmediated endothelial cell injury, Archives of Biochemistry and Biophysics, 294, 265–70. CHOE, S.Y., KIM, H.M. and YANG, K.H., 1984, Effects of butylated hydroxytoluene (BHT) on biliary excretion of xenobiotics and bile flow in rats, Drug and Chemical Toxicology, 7, 149–65. CONNING, D.M. and PHILLIPS, J.C., 1986, Comparative metabolism of BHA, BHT and other phenolic antioxidants and its toxicological relevance, Food and Chemical Toxicology, 24, 1145–8.

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CURRAN, P.G. and DEGROOT, L.J., 1991, The effect of hepatic enzyme-inducing drugs on thyroid hormones and the thyroid gland, Endocrine Reviews, 12, 135–50. GOPINATH, C., 1991, Causes and interpretation of thyroid tumours in rodent bioassays, in Current Issues in Drug Development, Eynsham, Oxford: HRC Information Press Inc. GRAHAM, M.J., WILSON, S.A., WINHAM, M.A., SPENCER, A.J., REES, J.A., OLD, S.L. and BONNER, F.W., 1994, Lack of peroxisome proliferation in marmoset liver following treatment with ciprofibrate for 3 years, Fundamental and Applied Toxicology, 22, 58–64. GRASSO, P. et al., 1991, Role of persistent, non-genotoxic tissue damage in rodent cancer and relevance to humans, Annual Review of Pharmacology and Toxicology, 31, 253–87. GREGUS, Z. and KLASSEN, C.D., 1988, Effect of butylated hydroxyanisole on hepatic glucuronidation and biliary excretion of drugs in mice, Journal of Pharmacy and Pharmacology, 40, 237–42. HAWKINS, J.M., JONES, W.E., BONNER, F.W. and GIBSON, G.G., 1987, The effect of peroxisome proliferators on microsomal, peroxisomal, and mitochondrial enzyme activities in the liver and kidney, Drug Metabolism Reviews, 18, 441–515. HILL, R.N., ERDREICH, L.S., PAYNTER, O.E., ROBERTS, P.A., ROSENTHAL, S.L. and WILKINSON, C.F., 1989, Thyroid follicular cell carcinogenesis, Fundamental and Applied Toxicology, 12, 629–97. ITO, N., FUKUSHIMA, S. and TSUDA, H., 1986a, Carcinogenicity and modification of the carcinogenic response by BHA, BHT, and other antioxidants, CRC Critical Reviews in Toxicology, 15, 109–50. ITO, N., FUKUSHIMA, S., TAMANO, S., HIROSE, M. and HAGIWARA, A., 1986b, Dose response in butylated hydroxyanisole induction of forestomach carcinogenesis in F344 rats, Journal of the National Cancer Institute, 77, 1261– 5. JOHNSON, S., McKILLOP, D., MILLER, J. and SMITH, I.K., 1993, The effects of rat thyroid function of an hepatic microsomal enzyme inducer, Human & Experimental Toxicology, 12, 153–8. KEITH, Y., CANNING, P.M., FOSTER, J., LHUGUENOT, J.C. and ELCOMBE, C. R., 1991, Peroxisome proliferation due to di-(2-ethylhexyl) adipate, 2-ethyl hexanol and 2-ethyl hexanoic acid, Archives of Toxicology, 66, 321–6. KRAUTSCHICK, I., KRUGMANN, J. and NEUENFELD, M., 1995, The effect of per-oxides on the vascular endothelium of isolated pig aorta in vitro, Experimental and Toxicologic Pathology, 47, 51–61. LAKE, B.G., EVANS, J.G., GRAY, T.J.B., KÖRÖSI, S.A. and NORTH, C.J., 1989, Comparative studies on nafenopin-induced hepatic peroxisome proliferation in the rat, Syrian hamster, guinea pig and marmoset, Toxicology and Applied Pharmacology, 99, 148–60. LIANG, H., MORIN, O. and BURGER, A.G., 1993, Effect of the antioxidant TK 12627 (Irganox) on monodeiodination and on the levels of messenger ribonucleic acid of 5′-deiodinase type I and spot 14, Acta Endocrinologica, 128, 451–8.

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LOCK, E.A., MITCHELL, A.M. and ELCOMBE, C.R., 1989, Biochemical mechanisms of induction of hepatic peroxisome proliferation, Annual Review of Pharmacology and Toxicology, 29, 145–63. McCLAIN, R.M., 1989, The significance of hepatic microsomal enzyme induction and altered thyroid function in rats: implications for thyroid gland neoplasia. Toxicologic Pathology, 17, 294–306. MITCHELL, A.M., LHUGUENOT, J.C., BRIDGES, J.W. and ELCOMBE, C.R., 1985, Identification of the proximate peroxisome proliferator(s) derived from di(2-ethylhexyl) phthalate, Toxicology and Applied Pharmacology, 80, 23–32. MUAKKASSAH-KELLY, S.F., KRINKE, A.-L., MALINOWSKI, W., STÄUBLI, W., BENTLEY, P., WAECHTER, F., JUGE-AUBRY, C. and BURGER, A.G., 1991, The effect of short term feeding of the antioxidant triethyleneglycol-bis-3 (3-tert-butyl-4-hydroxy-5-methyl)propionate on serum thyrotropin and thyroid hormones in the male rat, Toxicology and Applied Pharmacology, 107, 129– 40. PAYNTER, O.E., BURIN, G.J., JAEGER, R.B. and GREGORIO, C.A., 1988, Goitrogens and thyroid follicular cell neoplasia: evidence for a threshold process. Regulatory Toxicology and Pharmacology, 8, 102–19. PERCHELLET, J.-P. and PERCHELLET, E.M., 1989, Antioxidants and multistage carcinogenesis in mouse skin, Free Radical Biology and Medicine, 1, 377–408. PRICE, S.C., HINTON, R.H., MITCHELL, F.E., HALL, D.E., GRASSO, P., BLANAERD, G.F. and BRIDGES, J.W., 1986, Time and dose study on the response of rats to the hypolipidaemic drug fenofibrate, Toxicology, 41, 169– 91. PROCHASKA, H.J. and TALALAY, P., 1988, Regulatory mechanisms of monofunctional and bifunctional anticarcinogenic enzyme inducers in murine liver, Cancer Research, 48, 4776–82. RODRIGUES, A.D., FERNANDEZ, D., NOSARZEWSKI, M.A., PIERCE, W.M. JR. and PROUGH, R.A., 1991, Inhibition of hepatic microsomal cytochrome P-450 dependent monooxygenation activity by the antioxidant 3-tert-butyl-4hydroxyanisole, Chemical Research in Toxicology, 4, 281–9. SØNDERGAARD, D. and OLSEN, P., 1982, The effect of butylated hydroxytoluene (BHT) on the rat thyroid, Toxicology Letters, 10, 239–44. THOMAS, G.A. and WILLIAMS, E.D., 1991, Evidence for and possiblemechanisms of non-genotoxic carcinogenesis in the rodent thyroid, Mutation Research, 248, 357–70. VERHAGEN, H., FURNÉE, C., SCHUTTE, B., HERMANS, R.J.J., BOSMAN, F.T., BLIJHAM, G.H., TEN HOOR, F., HENDERSON, P.T. and KLEINJANS, J.C. S., 1989, Butylated hydroxyanisole-induced alterations in cell kinetic parameters in rat forestomach in relation to its oxidative cytochrome P-450mediated metabolism, Carcinogenesis, 10, 1947–51. VERHAGEN, H., SCHILDERMAN, P.A.E.L. and KLEINJANS, J.C.S., 1991, Butylated hydroxyanisole in perspective, Chemico-Biological Interactions, 80, 109– 34. VISSER, T.J., KAPTEIN, E., VAN TOOR H., VAN RAAIJ, J.A.G.M., VAN DEN BERG, K.J., TJONG TJIN JOE, C., VAN ENGELEN, J.G.M. and BROUWER A., 1993, Glucuronidation of thyroid hormone in rat liver: effects of in vivo

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treatment with microsomal enzyme inducers and in vitro assay conditions, Endocrinology, 133, 2177–86. WESTERMARK, K., KARLSSON, F.A. and WESTERMARK, B., 1985, Thyrotropin modulates EGF receptor function in porcine thyroid follicle cells, Molecular and Cellular Endocrinology, 40, 17–23. WILLIAMS, G.M., McQUEEN, C.A. and TONG, C., 1990a, Toxicity studies of butylated hydroxyanisole and butylated hydroxytoluene. I. Genetic and cellular effects, Food and Chemical Toxicology, 28, 793–8. WILLIAMS, G.M., WANG, C.X. and IATROPOULOS, M.J., 1990b, Toxicity studies of butylated hydroxyanisole and butylated hydroxytoluene, II. Chronic feeding studies. Food and Chemical Toxicology, 28, 799–806. ZBINDEN, G., 1987, Assessment of hyperplastic and neoplastic lesions of the thyroid gland, Trends in Pharmacological Sciences, 8, 11–14.

24 Toxicology of Surfactants: Molecular, Mechanistic and Regulatory Aspects WALTER STERZEL Henkel KGaA, Düsseldorf

Introduction The vast distribution of surfactants in various products in everyday use requires that the unwanted effects as well as the desired properties are known in order to recognize possible risks and prevent any damage to the health of humans. In order to understand the effects of surfactants on the organism, their most important biochemical effects, which depend on the interaction of surface active agents with basic biological structures like membranes, proteins and enzymes, are discussed. Following this discussion the local effects of surfactants are described. Local in this sense are all the effects encountered directly at the point of contact with the outer surfaces of the body, such as skin and mucous membrane irritation as well as allergies arising from skin contact. After this section the toxicokinetic properties of surfactants, providing information about type and extent of absorption by organisms, metabolic pathways and their elimination are discussed. The section on systemic effects deals, in contrast to local effects, with reactions arising after the substance has entered the organism by swallowing, skin penetration or inhalation. Due to their technical and economic importance, surfactants have been used extensively for decades. This resulted in an abundance of scientific publications concerning their effects on organisms. As a complete review would exceed the scope of this contribution, the focus will centre on a description of exemplary data which are important for the evaluation of the safety of surfactants. Biochemical properties of surfactants Surfactants come into immediate contact with the body during cleaning of the skin and act on the skin cells directly. When surfactants are swallowed unin tentionally, tissue damage is also possible. The question of the effects on the cells and cell components like membranes, proteins and enzymes is therefore also important from a toxicological point of view.

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Interactions with membranes Due to their ability to absorb at interfaces, surfactants can interact with biological membranes. This interaction depends on the concentration of the surfactant and can be described in the following sequence (Helenius and Simons, 1975). In the first instance the monomeric surfactant molecule adsorbs onto the membrane. For a low surfactant/membrane ratio this changes the permeability of the membrane and leads to cell lysis at higher concentrations. At even higher surfactant concentrations, the lamellar structure of the membrane is lost and it is solubilized. A further increase in surfactant concentration results in the separation of the phospholipids from the protein. This allows surfactant molecules to adsorb on previously hidden regions of the protein molecule. For the solubilization of integral membrane proteins the formation of micelle/protein complexes seems to be a prerequisite. A significant solubilization of these proteins is possible only if the critical micelle concentration cM is exceeded. This is indicated by the fact that the microsomal membrane bound enzyme arylsulphatase-C could only be extracted from the membrane with retention of the biological activity after micelles were formed (Chang et al., 1985). As a consequence of these interactions, surfactants are able to influence the metabolism of membrane components (DeLeo, 1989). This has been demonstrated by studies on the pathophysiology of surfactant-mediated skin irritation. In vitro cultured corneocytes showed an increased release of cholin metabolites after incubation with anionic surfactants. This effect was less pronounced after treatment with nonionic surfactants. In conclusion, these investigations demonstrated that the release of metabolites is correlated with the irritation potential of surfactants. Interactions with proteins Depending on the structure of the surfactant the interactions with proteins are based on polar or hydrophobic interactions. The binding of surfactants to protein molecules is a function of the concentration of free surfactant in equilibrium with the protein. The binding is affected by the pH, temperature and ionic strength of the solution. These factors can lead to conformational changes of proteins and thereby increase or decrease the number of available binding sites. Natural bovine albumin, for example, has 10 binding sites for decyl glucoside at 10°C and 13 at 25°C (Wasylewski and Kozik, 1979). According to a theory developed by Jones (1975), surfactants adsorb onto proteins in multiple equilibria. Only a few surfactant molecules (<10) are bound at high affinity binding sites during the first part of the adsorption process. In this process, no conformational changes of the protein are induced. A few water soluble proteins like bovine serum albumin, serum high-density lipoprotein, β-lactoglobulin and

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pyruvate oxidase can form this type of binding (Schwuger and Bartnik, 1980). Binding of more surfactant molecules leads to conformational changes in the protein. It is obvious that conformational changes allow binding of further surfactant molecules on hydrophobic regions which were previously not exposed. According to their different chemical structure (e.g. anionic, cationic, amphoteric or nonionic) surfactants differ significantly in their ability to carry out cooperative binding and therefore they differ in their biological activity. Anionic surfactants form adsorption complexes with proteins due to polar and hydrophobic interactions. Polar interactions between the negatively charged hydrophilic group of the surfactant and the positively charged groups of the protein molecule are the precondition for the formation of hydrophobic associations between surfactant molecule and protein molecule (Garcia-Dominguez, 1977; Schwuger and Bartnik, 1980). In the case of dodecylsulphate and tetradecylsulphate the binding results in denaturation of the proteins (Makino et al., 1973). Cationic surfactants can interact by polar and hydrophobic binding as well. Polar interactions result in electrostatic bonds between the negatively charged groups of the protein molecule and the positively charged surfactant molecule. For example, the enzyme, glucose oxidase, is deactivated by hexadecyl trimethyl ammonium bromide through formation of an ion pair between the cationic surfactant and the anionic amino acid side chain of the enzyme molecule (Tsuge, 1984). Nonionic or amphoteric surfactants and proteins show either no interaction at all or interactions that are extremely weak and normally close to the limits of sensitivity of the analytical methods used. For this reason, nonionic surfactants will not dissolve sparingly soluble proteins, denature proteins, or contribute to a swelling of the epidermis. Figure 24.1 shows the solubility of the protein zein, which is almost insoluble in water, and is more or less solubilized by sodium dodecyl sulphate and alkyl ethyleneglycol ether sulphates, while the nonionic ethoxylated nonylphenol is ineffective (Schwuger and Bartnik, 1980). A further reason for the poor interactions between nonionic surfactants and proteins could be that the concentration necessary for cooperative binding with the protein is not attained with nonionic surfactants due to their low critical micelle concentration cM (Makino et al., 1973). An important consequence of the interactions between anionic surfactants and proteins is the swelling of the stratum corneum of the skin. Hydrophobic interactions between surfactant chains and the protein result in pendant ionic head groups and subsequently in swelling because of electrostatic repulsion between them. As the substrate matrix expands and the tertiary structure is disrupted, hydration occurs which leads to swelling (Blake-Haskins, 1986).

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Figure 24.1 Zein solubility cz of saturated solutions as a function of surfactant concentration. Zein concentration, 50 g lit−1, mixing period, 2 h, temperature, 40°C ( , sodium dodecyl sulphate; , alkyl ether sulphate (2EO); , nonyl phenol ethoxylate (9EO)).

Interactions with enzymes Surfactants which are capable of massive cooperative binding, such as many anionic and cationic surfactants, induce conformational changes in the protein molecule which in general lead to loss of biological activity. The following mechanisms of enzyme inactivation by surfactants have to be considered (Ne’eman et al., 1971): 1. Disruption of the quaternary structure of the enzyme when the enzyme protein consists of several subunits.

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2. Induction of conformational changes in the tertiary or secondary structure of the enzyme protein. 3. In the case of membrane-bound enzymes, separation of the enzyme protein from essential membrane lipids. 4. Binding at active sites of the enzyme. While the effect of cationic surfactants on membranes is comparable to that of anionic surfactants, many proteins are obviously more resistant towards the denaturing activity of .cationic surfactants (Nozaki et al., 1974). Binding of tetradecyl trimethyl ammonium chloride onto bovine serum albumin and other proteins is comparable to that of sodium dodecyl sulphate. However, the cooperative binding with subsequent denaturation requires a ten-fold higher concentration of cationic surfactant. The saturation of the surfactant/protein complex is prevented by the competing formation of surfactant micelles. Contrary to the irreversibly denaturing effect of sodium dodecyl sulphate, the effect of some cationic surfactants on proteins is reversible (Nakaya et al., 1971). Local effects Skin compatibility The damaging effects of surfactants on skin manifest themselves in dryness, roughness and scaling. In addition, symptoms of inflammation (reddening, swelling) can develop, which can result, in severe cases, in complete destruction of the tissue. All these symptoms are a result of the described biochemical properties of surfactants. The skin is defatted by the more or less pronounced property of the surfactants to emulsify lipids and thus partially or completely removing the surface film of lipids. This leads to a disturbance of the barrier function of the skin resulting in increased permeability for chemical substances and a loss of water. Anionic surfactants can cause swelling of the skin. As a result, they facilitate the transport of substances to lower layers where inflammation reactions can be induced (Scholz, 1967). The reaction of surfactants with proteins dissolves proteins out of the skin and leads to their denaturation. These changes in the matrix material have an effect on the resistance of the skin (Götte, 1967) and, along with degreasing and drying, are an additional cause of an increase in skin roughness (Imokawa, 1975). The majority of the knowledge about skin compatibility of surfactants originates from studies with experimental animals, preferably rabbits. Furthermore, it is possible to evaluate new substances directly on human skin after careful exclusion of unreasonable risks. A critical overview of different test methods is given by Kästner (1980). In this context, the

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problem of labelling chemical products with respect to their toxicological properties has to be addressed. Different national or international regulations, e.g. the EG directive 67/548 within the European Community, dictate that these products are labelled ‘irritating’ or ‘corrosive’ whenever exactly defined effects are observed in appropriate tests. With consumer protection in mind, exceedingly stringent test procedures have been established. These conditions frequently result in an unfavourable classification especially for surfactants. When interpreting data from such studies, it is important to consider that unrealistic conditions of exposure were involved. Since anionic surfactants are the class with the greatest economic importance, they are the best studied. No general statement is possible with regard to a classification of the various groups of anionic surfactants in order of their skin compatibility, since within each class of substances significant differences exist in their effect on skin depending on the respective structure. Opdyke et al. (1965), for example, found a decrease in the skin irritation potential of different alkyl ether sulphates with increasing level of ethoxylation. The effect of the alkyl chain length of anionic surfactants was examined in different test models for soaps, alkyl sulphates, alkyl sulphonates, alkylbenzene sulphonates as well as alphaolefin sulphonates (Kästner, 1980). As shown in Table 24.1, it could be established in all cases that compounds with a saturated side chain of 10– 12 C atoms exert the largest effect, or rather, have the highest potential for damage. When the results of skin compatibility tests for the most important classes of anionic surfactants are summarized, it becomes evident that the undiluted products have to be regarded as strongly irritating substances. Even at concentrations of 10 per cent moderate to strong effects have to be expected. However, at concentrations less than 1 per cent, which is the range corresponding to typical use levels in detergents, only minimal irritation is observed. Nonionic surfactants have a good skin compatibility at normal use levels. Although studies with alcohol ethoxylates were reported in which a strong irritant effect was observed (Grupp et al., 1960), these studies used concentrations far above the usual exposure levels of consumers. Independent of their structure, cationic surfactants cause severe skin damage in high concentrations, while typical application levels are generally tolerated well. Mucous membrane compatibility When talking about mucous membrane compatibility one has to consider not only the mucous membranes of the eye. In addition, the mucous membranes in the mouth, upper and lower gastrointestinal tract as well as the urogenital tract have to be considered. In general, the effects of

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Table 24.1 Structure/activity relationships of anionic surfactants

a

Test model: A=epicutaneous, mouse; B=intracutaneous, mouse; C=epicutaneous, man; D=roughness of skin; E=swelling of collagen in vitro; F=denaturation of protein in vitro. b Number of carbon atoms in the alkyl chain.

surfactants on mucous membranes are based on the same biochemical mechanisms that are described in the chapter on skin compatibility. Special characteristics in the fine structure of mucous membranes, like the absence of keratin, result in a significantly higher sensitivity of these tissues towards chemical substances. Irritating materials affecting the eye cause reddening through increased blood flow in the conjunctivae with enlargement of the blood vessels. This can finally lead to the destruction of the cell walls accompanied by bleeding. Depending on the severity of the effects, a more or less pronounced swelling or reflex-induced closure of the eyelid will occur, followed by tearing and secretion. If the degree of irritation is low, epithelium damage develops on the cornea which can be visualized only with special techniques (staining, slit lamp microscope) and which is generally reversible. In severe cases the effects result in irreversible clouding of the cornea and therefore lead to an impairment of the eyesight. The classical method for the evaluation of mucous membrane compatibility of chemicals is the so-called Draize test on the rabbit eye (Draize et al., 1944). A structure/activity relationship with respect to the length of the respective alkyl chains of anionic surfactants can, as for the skin compatibility, also be observed for the mucous membrane compatibility (Kästner, 1980). According to this, the maximum irritation occurs at chain lengths of C10−14. for n-alkyl sulphates as well as for nalkyl sulphonates. Although the irritation potential of the different surfactant classes extends over a large range, it can be concluded that the mucous membrane compatibility decreases in the following order:

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nonionic>anionic>cationic surfactants (Draize and Kelley, 1952; Hazleton, 1952; Grant, 1962). Sensitization Aside from acute irritation, chemical substances can cause allergies after contact with the skin or a mucous membrane. The development of an allergy is dependent on certain preconditions. An essential factor is the individual disposition which is predominantly genetically determined. An additional important point is the extent of damage to the tissue at the point of contact of the chemical substance (inflammation), which promotes sensitization. In addition, the sensitization potential of a substance is of decisive importance. For products with low molecular weights, this potential is dependent on their chemical properties. Small molecules are by themselves not able to trigger a reaction of the immune system. They become immunologically active only after binding to endogeneous proteins. Since the majority of the surfactants can only form weak and reversible bindings via hydrophobic and electrostatic interactions, this prerequisite is not fulfilled. Once the organism is sensitized towards a certain chemical, renewed contact with trace amounts of this material can provoke allergic reactions, which especially affect the skin and respiratory tract. Typical symptoms are itching, eczema, exanthema, rhinitis and bronchial asthma. Anionic surfactants and surfactant containing products were tested for sensitizing properties by numerous laboratories (Götte, 1967; Kästner, 1980; Siwak et al., 1982) without detecting any significant increase in risk. The same holds true for nonionic surfactants (Siwak et al., 1982). Some cationic surfactants, which are able to form stable complexes by the formation of ion pairs with anionic groups of proteins, proved to be allergenic (Schallreuter and Wood, 1986). Toxicokinetics Percutaneous absorption The most important exposure of humans occurs through the skin with the use of cosmetics and toiletries. The skin comes in contact with surfactants also during dishwashing or when washing hands. Since these products are used over a long period of time, possible long-term effects must be evaluated. Measurement of percutaneous absorption of surfactants is important because it provides data for the toxicologist concerning the amount of surfactants which could enter the body through the skin in the most unfavourable case. Together with other toxicological information,

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this allows a realistic evaluation of the risk when these compounds are used. Due to their economic importance, most studies have been carried out with anionic surfactants. Fewer studies exist for the other classes of surfactants. In vitro measurements of the percutaneous absorption of sodium dodecyl sulphate indicated a low absorption value for rat skin as well as for human skin (Blank and Gould, 1961; Embery and Dugard, 1969; Howes, 1975). The low cutaneous absorption of sodium dodecyl sulphate could also be confirmed in experiments with rats (Greb, 1980). After application of a 0.7 per cent aqueous solution of sodium dodecyl sulphate (contact time 15 min), a cutaneous absorption of 0.26 µg cm−2 within 24 h was measured (Howes,1975). In summarizing the results of the available studies, one can conclude that only small amounts of surfactants are resorbed through the intact skin. Since human skin in general is less permeable to chemicals (Rice, 1977; Wester and Maibach, 1982), the amounts of surfactants absorbed cutaneously in everyday use are probably even smaller. If the epidermis is removed completely or partially, e.g. damaged skin, the degree of absorption can increase substantially (Scala et al., 1968). In vitro studies demonstrated that cationic surfactants are absorbed by the skin to a much lesser extent than anionic surfactants (Scala et al., 1968; Geisler, 1976; Faucher et al., 1979). The degree of percutaneous absorption is generally larger for nonionic surfactants than for anionic or cationic surfactants. Studies on the percutaneous absorption of alkyl polyethyleneglycol ethers of the structure C12-(CH2-CH2-O)3H, C12-(CH2-CH2-O)6H, C12-(CH2-CH2-O)10H, and C15-(CH2-CH2-O)3H were performed under conditions of use (Black and Howes, 1979). The aqueous solutions of the applied surfactants were washed off after a contact time with the skin of 15 min. Under these conditions, the penetration of the alkyl polyethyleneglycol ethers was greater than the penetration of the analogous alcohol sulphates or alcohol ether sulphates. The penetration increased with increasing length of the carbon chain. Percutaneous absorption decreases for an ethylene oxide content of 6 moles or more in the ethoxylate moiety. Intestinal absorption, metabolism and excretion The ingestion of surfactants is possible e.g. through the use of surfactantcontaining toothpaste, through residues from dishwashing detergents and through traces of surfactants in potable water. Anionic surfactants are resorbed well in the intestine (Michael, 1968; Black and Howes, 1980; Bartnik and Künstler, 1987). After absorption, a part of this is excreted together with bile in the faeces and is subject to a enterohepatic cycle. The majority of the absorbed surfactant is metabolized in the liver

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and the respective metabolites are eliminated in the urine. The metabolic degradation of the linear alkyl chain is performed by -oxidation followed by β-oxidation. The ether-linkage in the ethoxylate portion of sulphated alcohol ethoxylates seems to be resistant to metabolism. Linear alkylbenzene sulphonates and branched alkylbenzene sulphonates are metabolized to short chain sulphophenyl carboxylic acids. N-alkyl sulphates are metabolized by -oxidation of the hydrophobic end followed by β-oxidation. Butyric acid-4-sulphate and acetic acid-2-sulphate are the end products, which are then further converted in small amounts nonenzymatically to sulphate and -butyrolactone (Ottery et al., 1970). Studies by Taylor et al. (1978) demonstrated that alkyl sulphonates are degraded via the same pathway as alkyl sulphates. Cationic surfactants can be assumed to be resorbed in the intestine only to a small extent. This was confirmed in a study with trimethyl cetyl ammonium bromide (Isomaa, 1975; Isomaa et al., 1976). Due to the low level of resorbed surfactant, an unquestionable identification of the metabolites was not possible. Parts of absorbed cationic surfactants were, as found for anionic surfactants, excreted together with bile in the faeces and to a lesser degree with the urine. Nonionic surfactants are resorbed to a large degree in the intestine (Drotman, 1980). A significant part of the material is eliminated with the bile. Cleavage of the ether linkage is obviously possible. Homologous ethyleneglycol ethers are probably generated as metabolites along with the corresponding carboxylic acids which are formed through oxidation of the terminal hydroxymethyl group (Drotman, 1980). The sorbitan fatty esters, which are often used as emulsifiers, and the ethoxylated fatty acid esters are hydrolysed in the gastrointestinal tract after oral administration through cleavage of the ester bond. While the resulting fatty acid is treated metabolically like a natural fatty acid, the polyol component of the sorbitan fatty acid is absorbed in the intestine, but is not further oxidized and is eliminated predominantly with the urine (Elworthy and Treon, 1967). Systemic effects Talking about systemic effects means, in contrast to local effects, the description of reactions arising after the substance has entered the organism after swallowing, skin penetration or inhalation. For surfactants, resorption through the skin has to be considered in particular. As described in the previous section, it is relatively small. But for products that frequently come into close contact with the skin, either unintentionally or due to their intended use, the resorption of very small amounts over a long period of time cannot be prevented.

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Acute toxicity In general, the acute oral toxicity of surfactants is low. The LD50 values typically range between several hundred and several thousand mg kg−1 of bodyweight. This is of the same order of magnitude as for table salt (Swisher, 1968). The most important effects are damage to the mucous membranes of the gastrointestinal tract. High doses induce vomiting and diarrhea (Weaver and Griffith, 1969). Surfactants exhibit significantly higher toxicity when the gastrointestinal tract is by-passed through intravenous injections. Even at very low concentrations, the interaction with the membrane of erythrocytes leads to their destruction. Inhalation of surfactant-containing dusts or aerosols in higher concentrations leads to disturbances of the lung function (Coate et al., 1978). This effect can be attributed to interactions with the surface active film that lines the vesicles of the lung (Kissler et al., 1981). As with local compatibility, there are also pronounced structure/activity relationships for acute toxicity. Gale (1953) has investigated the acute toxicity of sodium alkyl sulphates from C8 to C18 and found the strongest effect for C12 sulphate. The anaesthetic properties of certain alcohol ethoxylates which can be observed after intravenous application as well as after application to the skin or the mucous membranes are remarkable. Ethoxylates of unbranched primary alcohols with 9 ethylene oxide units were found to exhibit local anaesthetic properties starting with an alkyl chain of C8. The activity increases with increasing chain length (Zipf and Dittmann, 1964). Chronic toxicity In order to exclude any adverse effects arising from the repeated exposure against small amounts of surfactants over a prolonged period of time, representatives of all important classes of surfactants were examined for chronic toxic effects. In these tests, dosages of several thousands ppm were administered over a period of up to 2 years. No observable effects were detected with linear alkylbenzene sulphonates in 2 year studies with rats using concentrations up to 0.5 per cent (feed) or 0.1 per cent (drinking water) (Buehler et al., 1971). A sodium alkyl sulphate with an average chain length of C12 was tolerated by rats up to 1 per cent in the feed for 1 year without any remarkable side effects (Fitzhugh and Nelson, 1948). C14 −16 α-olefin sulphonates were applied over 2 years in a feeding study in dosages up to 0.5 per cent without causing any remarkable effect (Hunter and Benson, 1976). Analogous studies were reported for alcohol ethoxylates and alkylphenol sulphates, which revealed no toxic symptoms at doses up to 0.1 per cent and 1.4 per cent, respectively (Larson et al., 1963; Siwak et al., 1982). Studies on cationic surfactants reported a noobservable-effect-level of 0.25 per cent (Coulston et al., 1961). In all these

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long-term studies, the dosages that were tolerated without damage were in the range of several thousand ppm, indicating large margins of safety. This was confirmed by Hunter and Benson (1976), who calculated for a relevant example that the respective dosage lies at least by a factor of 1000 over the estimated maximum daily exposure level of humans. Besides these data from animal experiments, a series of studies exists in which volunteers ingested considerable amounts of anionic or nonionic surfactants over several weeks, without any noticeable severe adverse effects (Swisher, 1968). Mutagenicity Mutagenicity is the induction of irreversible changes in genetic material. If normal cells (somatic cells) are the target, malformation results in the developing organism. In case of the mature organism, it can lead to tumour formation. If germ cells are the target, the danger exists that the genetic defect will be passed on to the offspring. All classes of surfactants have been evaluated in numerous test systems. The collected data allow the conclusion that surfactants pose no considerable risk of genetic damage (Yam et al., 1984; Fowler, 1988; Oba and Takei, 1992). Carcinogenicity Due to the widespread use and contact with surfactants the question of irreversible damage has to be raised in addition to the problem of other chronic effects. The following compounds were evaluated for carcinogenicity after administration in the drinking water or feed: alkylbenzene sulphonate (Buehler et al., 1971), alkyl sulphates (Fitzhugh and Nelson, 1948), α-olefin sulphonates (Hunter and Benson, 1976), secalkane sulphonate (Quack and Rend, 1976), alcohol ether sulphates (Tusing et al., 1962; Siwak et al., 1982), alcohol ethoxylates (Siwak et al., 1982) and alkylphenol ethoxylates (Larson et al., 1963; Smyth and Calandra, 1969). None of these experiments provided any indication of increased risk of cancer after oral ingestion of surfactants. The question of possible carcinogenic effects of surfactants on the skin has also been studied extensively. Summaries exist by Oba and Takei (1992) and Siwak et al. (1982). Embryotoxicity The effects of substances on the organism during pregnancy can lead to delayed development or death of the embryo or malformation. Studies with the following surfactants revealed no indications of embryotoxic activity: alcohol ethoxylates (Nomura et al., 1980), α-olefin sulphonates (Palmer et

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al., 1975), alcohol ether sulphates and linear alkylbenzene sulphonates (Nolen et al., 1975). Concerns which started with the publication in 1969 (Mikami et al., 1969) that surfactants had caused malformations in animal studies could not be reproduced (Oba and Takei, 1980). The findings of Mikami et al. (1969) were interpreted to be a result of methodical inadequacies and misinterpretations (Charlesworth, 1976). Summary Due to their physico-chemical properties, surfactants are capable of reacting with biological membranes, proteins and enzymes. Most of their toxicological properties can be traced back to these interactions. During the application of surfactant-containing products, the most important aspect of consumer safety is local compatibility. No indications of systemic, chronic or irreversible damage could be found. Estimates of the amounts of orally ingested surfactants typically encountered were reviewed by several authors. Based on these estimates, a total daily intake of surfactants in the range of 0.3–3 mg per person was calculated by Swisher (1968). Due to the low rate of percutaneous absorption exposure through the skin can be neglected. If the above mentioned highest conceivable daily intake is compared with the dosage that was tolerated without adverse effects in studies concerning systemic effects, it becomes quite clear that these amounts can be regarded as harmless. In conclusion, it can be stated that the use of surfactants does not pose a health risk for humans. References BARTNIK, F.G. and KÜNSTLER, K., 1987, Biological effects, toxicology and human safety, in Falbe, J. (Ed.). Surfactants in Consumer Products, Theory, Technology and Application, p. 475, Heidelberg: Springer Verlag. BLACK, J.G. and HOWES, D., 1979, J. Soc. Cosmet. Chem., 30, 157. BLACK, J.G. and HOWES, D., 1980, Absorption, metabolism, and excretion of anionic surfactants, in Gloxhuber, C. (Ed.) Anionic Surfactants, Biochemistry, Toxicology, Dermatology; Surfactant Sci. Ser., Vol 10, p. 51, New York, Basel: Marcel Dekker. BLAKE-HASKINS, J.C., 1986, J. Soc. Cosmet. Chem., 37, 199. BLANK, H.J. and GOULD, E., 1961, J. Invest. Dermatol, 37, 311. BUEHLER, E.V., NEWMAN E.A. and KING, W.R., 1971, Toxicol. Appl. Pharmacol, 18, 83. CHANG, P.L., AMEEN, M., LAFFERTY, K.I., VAREY, P.A., DAVIDSON, A.R. and DAVIDSON, R.G., 1985, Anal. Biochem., 144, 362. CHARLESWORTH, F.A., 1976, Food Cosmet. Toxicol, 14, 152. COATE, W.B., BUSEY, W.M., SCJOENFISCH, W.H., BROWN, N.M. and NEWMAN, E.A., 1978, Toxicol Appl. Pharmacol, 45, 477.

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COULSTON, F., DROBECK, H.P., MIELENS, Z.E. and GARVIN, P.J., 1961, Toxicol. Appl. Pharmacol., 3, 584. DELEO, V.A., 1989, J. Toxicol-Cut. Ocular Toxicol., 8, 227. DRAIZE, J.H., WOODWARD, G. and CALVERY, H., 1944, J. Pharmacol. Exp. Ther., 82, 377. DRAIZE, J.H. and KELLEY, E.A., 1952, Drug Cosmet. Ind., 71, 36. DROTMAN, R.B., 1980, Toxicol. Appl. Pharmacol, 52, 38. ELWORTHY, P.H. and TREON, J.F., 1967, Physiologic activity of nonionic surfactants, in Schick, M.J. (Ed.) Nonionic Surfactants, p. 923, New York, Basel: Marcel Dekker. EMBERY, G. and DUGARD, P.H., 1969, Br. J. Derm., 81 Supplement 4, 63. FAUCHER, J.A., GODDARD, E.D. and KULKARNI, R.D., 1979, J. Am. Oil Chem. Soc., 56, 776. FITZHUGH, O.G. and NELSON, A.A., 1948, Am. Pharm. Assoc., 37, 29. FOWLER, C., 1988, Toxic. in Vitro, 2, 65. GALE, L.E. and SCOTT, P.M., 1953, J. Am. Pharm. Assoc., Sci. Ed., 42, 283. GARCIA DOMINGUEZ, J., 1977, J. Cosmet. Chem., 28, 165. GEISLER, R.W., 1976, Toxicol. Appl. Pharmacol, 37, 98. GÖTTE, E., 1967, Tenside, 4, 209. GRANT, W.W., 1962, Toxicology of the Eye, p. 511, Springfield 111: Charles C. Thomas. GREB, W. and WINGEN, F., 1980, Seifen, Fette, Öle, Wachse, 106, 327. GRUPP, T.C., DICK, L.C. and OSER, M., 1960, Toxicol. Appl. Pharmacol., 2, 133. HAZLETON, L.W., 1952, Proc. Sci. Sect. Toilet Goods Ass., 17, 5. HELENIUS, A. and SIMONS, K., 1975, Biochim. Biophys. Acta, 414, 29. HOWES, D., 1975, J. Soc. Cosmet. Chem., 26, 47. HUNTER, B. and BENSON, H.G., 1976, Toxicology, 5, 359. IMOKAWA, G., SUMURA, K. and KATSUMI, M., 1975, J. Am. Oil Chem. Soc., 52, 484. ISOMAA, B., REUTER, J. and DJUPSUND, B.M., 1976, Arch. Toxicol, 35, 91. ISOMAA, B., 1975, Food Cosmet. Toxicol, 13, 231. JONES, M.N., 1975, Biochem. J., 151, 109. KÄSTNER, W., 1980, Local tolerance (animal tests): mucous membranes and skin, in Gloxhuber, C. (Ed.) Anionic Surfactants, Biochemistry, Toxicology, Dermatology; Surfactant Sci. Ser., Vol. 10, pp. 127, New York, Basel: Marcel Dekker. KISSLER, W., MORGENROTH, K. and WELLER, W., 1981, Prog. Resp. Res., 15, 121. LARSON, P.S., BORZELLECA, J.F., BOWMAN, E.R., CRAWFORD, E.M., SMITH, R.B.Jr., and HENNIGAR, G.R., 1963, Toxicol. Appl Pharmacol., 5, 782. MAKINO, S., REYNOLDS, J.A. and TANFORD, C., 1973, J. Biol. Chem., 248, 4926. MICHAEL, W.R., 1968, Toxicol. Appl Pharmacol., 12, 473. MIKAMI, Y., NAGAI, H., SAKAI, Y., FUKUSHIMA, S. and NISHINO, T., 1969, Cong. Anom. (Jap.), 9, 230.

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NAKAYA, K., YAMADA, K., ONOZAWA, M. and NAKAMURA, Y., l97l, Biochem. Biophys. Acta, 251, 7. NE’EMAN, Z., KAHANE, J. and RAZIN, S., 1971, Biochem. Biophys. Acta, 249, 169, 176. NOLEN, G.A., KLUSMAN, L.W., PATRICK, L.F. and GEIL, R.G., 1975, Toxicology, 4, 231. NOMURA, T., KIMURA, S., HATA, S., KANZAKI, T. and TANAKA, H., 1980, Life Sci., 26, 49. NOZAKI, Y., REYNOLDS, J.A. and TANFORD, C., 1974, J.Biol. Chem., 249, 4452. OBA, K. and TAKEI, R., 1992, Carcinogenic, mutagenic/genetic toxicity, and teratogenic properties, in Gloxhuber, C. and Künstler, K. (Eds) Anionic Surfactants, Biochemistry, Toxicology, Dermatology; Surfactant Sci. Ser., Vol. 43, p. 331, New York, Basel: Marcel Dekker. OPDYKE, D.L. and BURNETT, M.C., 1965, Proc. Sci. Sect. Toilet. Goods Assoc., 44, 3. OTTERY, J., OLAVESEN, A.A. and DODGSON, K.S., 1970, Life Sci., 9, 1335. PALMER, A.K., READSHAW, M.A. and NEUFF, A.M., 1975, Toxicology, 3, 107. QUACK, J.M. and REND, A.K., 1976, Fette-Seifen-Anstrichmittel, 78, 200. RICE, D.P., 1977, Appl. Pharmacol., 39, 377. SCALA, J., McOSTER, D.E. and RELLER, H.H., 1968, J. Invest. Dermatol., 50, 371. SCHALLREUTER, K. and WOOD, J.M., 1986, Biochem. Biophys. Res. Commun., 135, 221. SCHOLZ, J., 1967, Arch. Exp. Pathol. Pharmacol, 232, 241. SCHWUGER, M.J. and BARTNIK, F.G., 1980, Interactions of anionic surfactants with proteins, enzymes and membranes, in Gloxhuber, C. (Ed.) Anionic Surfactants, Biochemistry, Toxicology, Dermatology; Surfactant Sci. Ser., Vol. 10, p. 1, New York, Basel: Marcel Dekker. SIWAK, A., GOYER, M., PERWAK, J. and THAYER, P., 1982, in Mittal, K.L., Fendler, E.J. (Eds), Solution Behavior of Surfactants, Vol. I, p. 161, New York, London: Plenum Publishing Corp. SMYTH, J.F. and CALANDRA, J.C., 1969, Toxicol. Appl. Pharmacol., 14, 315. SWISHER, R.D., 1968, Arch. Environ. Hth, 17, 232. TAYLOR, A.J., POWELL, G.M., HOWES, D., BLACK, J.G. and OLAVESEN, A. H., 1978, Biochem. J., 174, 405. TSUGE, H., 1984, Agric. Biol. Chem., 48, 19. TUSING, T.W., PAYNTER, O.E., OPDYKE, D.L. and SNYDER, F.H., 1962, Toxicol. Appl. Pharmacol., 4, 402. WASYLESWSKI, Z. and KOZIK, A., 1979, Eur. J. Biochem., 95, 121. WEAVER, J.E. and GRIFFITH, J.F., 1969, Toxicol. Appl Pharmacol., 8, 214. WESTER, R.C. and MAIBACH, H.I., 1982, Occup. Ind. Dermatol, p. 201, Chicago: Year Book, Med. YAM, J., BOOMAN, K.A., BRODDLE, W., GEIGER, L., HEINZE, J.E., LIN, Y. J., MCCARTHY, K., REISS, S., SAWIN, V., SEDLAK, R.I., SLESINSKI, R.S. and WRIGHT, G.A., 1984, Food Chem. Toxicol., 22, 761. ZIPF, H.E. and DITTMANN, E.C., 1964, Arch. Exp. Pathol. Pharmacol, 247, 544.

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25 Low Dose of a Genotoxic Carcinogen does not ‘Cause’ Cancer; it Accelerates Spontaneous Carcinogenesis WERNER K.LUTZ University of Würzburg, Würzburg

Definitions of cancer risk The risk of cancer is normally expressed as a fraction of a population diagnosed with cancer within a specified period of time. In an animal bioassay for carcinogenicity, this period usually is 2 years; in cancer epidemiology, a life span of 65 years (0–64) is often used. The two periods can be considered equivalent with respect to the process of carcinogenesis: its rate in different species is inversely correlated with the natural life span and basal metabolism and the background cancer incidence in 2-year old rats or mice is very similar to the one seen in 65-year-old humans (Anisimov, 1989; Raabe, 1989; Tennant, 1993). For an individual, a cancer risk can only be 0 or 1, depending on whether the situation is analysed before or after the diagnosis of the tumour. The population-based expression of a cancer risk therefore is not easily visualized and does not take into account interindividual differences in susceptibility. In the following discussion, the dose-response relationship in chemical carcinogenesis is analysed in terms of an effect of a carcinogen on the individual tumour latency time (Kodell et al., 1980; Littlefield et al., 1980; Day, 1983; Gaylor, 1992). Together with the idea of background DNA damage responsible for what is considered ‘spontaneous’ tumour formation and including individual variability for the rate of this process, it will be shown that low doses of genotoxic carcinogens might accelerate the spontaneous process of carcinogenesis but are not expected to induce cancer ‘out of the blue’. Linear dose response for a DNA-reactive carcinogen The effect of a carcinogen at low dose is normally extrapolated from data obtained in 2-year bioassays. At the end of the 2-year treatment period, the surviving animals are killed and analysed for the presence of tumours. The

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fraction of tumour-bearing animals is then plotted against the dose, and lowdose effects are estimated from some model curve fitted to the data points. The shape of the dose-response curve at the low-dose end is strongly debated, especially for nongenotoxic carcinogens. For DNA-reactive carcinogens, it is widely accepted that there is no dose without effect, and a linear extrapolation is used (Lutz, 1990b). This is based on the idea that 1 molecule of a DNA-reactive carcinogen could form a dangerous DNA adduct in a critical gene and activate an oncogene or inactivate a tumour suppressor gene by mutation, if the adduct is not repaired before DNA replication. Table 25.1 shows the consequences of linear interpolation between the tumour incidence in the controls and in a dosed group of a bioassay for carcinogenicity. An organ-specific tumour incidence of 4 per cent in the control group (2/50) and 14 per cent (7/50) after treatment for 2 years at 10 mg kg−1 per day is assumed. With linear interpolation, a treatmentrelated increment in tumour incidence of 1 per cent would be calculated per mg kg−1 per day so that at 1 mg kg−1 per day, a 5 per cent total tumour incidence would be expected. In humans, increments in cancer risk in the per cent range would not be acceptable. A risk of 1 in 1 million lives might be considered negligible and the respective exposure could be regarded as ‘virtually safe.’ With the example given in Table 25.1 and upon linear interpolation, this ‘virtually safe’ dose would be calculated as 0.0001 mg kg−1 per day. What does it mean: ‘1 additional tumour in 1000 000 lives’? The fact that a cancer risk is only 10−6 cannot be a consolation for the affected individual. For this person, the cancer risk was 1. In the public opinion, therefore, an increase by one tumour case per one million lives is often interpreted to mean that one additional individual has got cancer who could otherwise have lived a much longer tumour-free life. For reasons explained below, this fear appears unfounded. Endogenous DNA damage; individual susceptibility Carcinogenesis is a multi-stage process based on the accumulation of a number of critical DNA-related changes, due to, for example, DNAcarcinogen adducts. Evidence of background DNA damage from endogenous and unavoidable substances is accumulating (Ames, 1989; Loeb, 1989; Lutz, 1990a). It is due to, for instance, electrophiles such as S-adenosylmethionine, epoxides, quinones, or aldehydes, or to reactive oxygen species. In addition, DNA is not a chemically stable molecule, it

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Table 25.1 Linear low-dose interpolation of a cancer risk based on a hypothetical 2year bioassay for carcinogenicity

Notes: Data in boldface; extrapolations in italics.

depurinates and deaminates spontaneously. Finally, DNA replication is not 100 per cent correct so that mutations cannot be avoided completely. The resulting background DNA damage is responsible for spontaneous mutations and for what is called spontaneous tumour formation. The process of carcinogenesis therefore has a non-zero rate even if exposure to exogenous DNA-reactive carcinogens could be avoided. The level of the background mutation rate is expected to show interindividual variability. It depends both upon genetic and life-style factors which govern, for instance, enzyme activities responsible for carcinogen metabolism or DNA repair (Harris, 1989). The rate of spontaneous carcinogenesis is further governed by the inherited and acquired presence of activated oncogenes or absence of tumour suppressor genes (Scrable et al., 1990). Therefore, each individual in a heterogeneous population is expected to have its own endogenous cancer risk expressed as an individual time-to-tumour or tumourfree lifetime. Exogenous DNA damage; acceleration of spontaneous carcinogenesis Exposure to an additional, exogenous DNA-reactive molecule adds to the background DNA damage, increases the probability of a mutation and accelerates the multi-stage process of carcinogenesis. At low doses of the exogenous carcinogen, the rate of the process is expected to be dominated by the background damage so that the exogenous factor cannot constitute a cancer risk independent of the spontaneous process. The acceleration must be dose dependent and might be related to the background rate operating in each individual. It is true, therefore, that even a few molecules of a DNA-reactive carcinogen can have an effect. However, this effect cannot be a tumour

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Figure 25.1 Schematic representation of the time course of tumour appearance in a group of individuals with large differences in susceptibility. Solid line: background process of spontaneous carcinogenesis; arrows: acceleration of the spontaneous process by exposure to an additional carcinogen.

‘out of the blue’, in an individual that would otherwise have a low cancer risk. It ‘only’ reduces the individual’s tumour-free lifetime. No cancer ‘out of the blue’ This interpretation does not contradict the understanding that a low dose of a carcinogen could increase the tumour incidence from 40000 to 40001 per 1000000 lives, in the example shown in Table 25.1. The connection between the two approaches is shown in Figure 25.1. The solid line shows the appearance of a spontaneous tumour in individuals of a group of 20 people. At the age of 65 years, 4 individuals have a tumour diagnosed. This is equivalent to a cumulative tumour incidence of 20 per cent. Exposure of this group to an additional exogenous carcinogen would result in some reduction in the tumour-free lifespan in all individuals. In the example shown in Figure 25.1, this shift would move one additional individual to an age of diagnosis <65 years (marked with *). The cumulative incidence 0– 64 now is 5 of 20, i.e. it has increased by one case, i.e., from 20 to 25 per cent. With decreasing dose, the length of the arrows shown in Figure 25.1 would also decrease. At low dose, and in an exposed population of 1 million, there could be one individual with a tumour now diagnosed at the age of 64.999 years who otherwise would have shown it only at 65.001

360 CARCINOGENESIS AT LOW DOSE

years. This is equivalent to an increase from 40000 to 40001 as a cumulative incidence 0–64, but it has a completely different meaning. It can now be excluded that the additional individual would have lived tumour free for 80 years in the absence of the exogenous carcinogen. Final remarks With the ideas presented, fear of cancer from low dose or from rare exposures can possibly be reduced. This does not mean that small cancer risks should be tolerated. Carcinogens in the environment, for instance, affect all of us; the tumour-free life span is reduced in the entire population. The model should be valid for tissues with a high spontaneous tumour incidence and an exponentially steep age dependence indicative of a multistage requirement of 5–6 steps. For cells that can be transformed in 2 or 3 steps or that are specifically sensitive in certain phases of the development (in utero or during childhood), the model has to be reconsidered. This might be necessary for tumours with incidence peaks at a young age (leukaemia, tumours of the lymphatic tissues, brain, testis). Nevertheless, the latter tumour types are rare in comparison with cancer of the old age so that the concept should hold for the majority of the tumours in humans. References AMES, B.N., 1989, Endogenous DNA damage as related to cancer and aging, Mutat. Res., 214, 41–6. ANISIMOV, V.N., 1989, Dependence of susceptibility to carcinogenesis on species life span, Arch. Geschwulstforsch., 59, 205–13. DAY, N.E., 1983, Time as a determinant of risk in cancer epidemiology: the role of multi-stage models, Cancer Surv., 2, 577–93. GAYLOR, D.W., 1992, Relationship between the shape of dose-response curves and background tumour rates, Regul. Toxicol. Pharmacol., 16, 2–9. HARRIS, C.C., 1989, Interindividual variation among humans in carcinogen metabolism, DNA adduct formation and DNA repair, Carcinogenesis, 10, 1563–6. KODELL, R.L., FARMER, J.H., LITTLEFIELD, N.A., FRITH, C.H., 1980, Analysis of life-shortening effects in female Balb/c mice fed 2-acetylaminofluorene, J. Environ. Pathol. Toxicol, 3 69–88. LITTLEFIELD, N.A., FARMER, J.H. and GAYLOR, D.W., 1980, Effects of dose and time in a long-term, low-dose carcinogenic study, J. Environ. Pathol. Toxicol., 3, 17–34. LOEB, L.A., 1989, Endogenous carcinogenesis: molecular oncology into the twentyfirst century—Presidential address, Cancer Res., 49, 5489–96.

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LUTZ, W.K., 1990a, Endogenous genotoxic agents and processes as a basis of spontaneous carcinogenesis, Mutat. Res., 238, 287–95. LUTZ, W.K., 1990b, Dose response relationship and low dose extrapolation in chemical carcinogenesis, Carcinogenesis, 11, 1243–7. RAABE, O.G., 1989, Scaling of fatal cancer risks from laboratory animals to man, Hlth Phys., 57 suppl. 1, 419–32. SCRABLE, H.J., SAPIENZA, C. and CAVENEE, W.K., 1990, Genetic and epigenetic losses of heterozygosity in cancer predisposition and progression, Adv. Cancer Res., 54, 25–62. TENNANT, R.W., 1993, Stratification of rodent carcinogenicity bioassay results to reflect relative human hazard, Mutat. Res., 286, 111–18.

26 Controversial Mechanistic and Regulatory Issues in Safety Assessment of Industrial Chemicals —an Industry Point of View HEINZ-PETER GELBKE BASF AG, Ludwigshafen

Introduction In the appropriate classification and risk assessment of industrial chemicals three main players are involved: the scientific community, regulatory authorities and the chemical industry. The rules of the game are given by the definitions of the classification criteria and by guidelines for the risk assessment process. These definitions and guidelines are sometimes very flexible and open to different interpretations but in other cases precisely defined. Mostly these rules have been set up by regulatory bodies, e.g. the EU or national authorities, but sometimes also by scientific committees like the MAK commission in Germany or by IARC. Looking now at these three main players, the scientific community will provide the data as the starting point for each individual chemical. Possible controversies centre around the question, how data gaps may be bridged by scientifically valid assumptions. This indeed may often be necessary, if a complete toxicological and mechanistic data base is not available. Nevertheless, a scientifically based consensus can often be achieved for this bridging process. On the other hand, the other two players—industry and regulatory authorities—often do not reach mutually agreed decisions, although both of them finally strive for the same target: protection of human health and the environment in an industrialized community. Controversial issues may sometimes stem from different approaches for bridging the scientific data gaps, but mostly they arise from political, economic or social aspects, which often are not outspoken. Just to give some examples: anticipated reaction of the society, possible emotions of the consumer, possible influences on the next election, availability of technical alternatives, different perceptions of the risk-benefit balance, overall economic situation of the community, different evaluations in other countries, impact on worldwide competitiveness, etc. In the following an industry point of view will be presented for some specific problems of today in the area of classification which is a more

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qualitative approach and for risk assessment which in addition has to take into account quantitative aspects. This will be discussed separately for toxicological effects with thresholds (‘classical’ organ toxicity, reproductive and developmental toxicity) and without thresholds (mutagenicity, carcinogenicity). Apart from this subdivision there is one general concern of industry, that is to appropriately take into account exposure. Exposure Every classification and especially risk assessment decision should not only be based qualitatively on the toxicological profile, but it should also take into account quantitatively the toxicological dose-response relationship as compared to human exposure. As a first rough approximation the different exposure profiles may be grouped into four main categories: 1. Exposure during chemical production Many high production volume chemicals are used mainly or even exclusively as intermediates within the chemical industry. Although large amounts of these materials may be produced or processed within only a few facilities, exposure is often quite low and can be controlled or reduced by technical means. In addition there are many specific features enabling an efficient exposure control, such as: a trained workforce, site-specific and personal protection devices, stringent surveillance of workforce and work procedures, medical programmes tailored to the specificities of the work place, well defined exposures at the specific work sites, the exposed population is well known and limited, specified exposure duration, relatively homogeneous age and better health status of the workforce as compared to the general population, etc. 2. Exposure of the downstream user during industrial/ manufacturing applications In principle, the exposure scenarios may be quite similar to those of chemical production, since many of the features described above also relate to or may be implemented at smaller workshops of the downstream user. Unfortunately, in reality often quite high exposures prevail in small workshops, possibly due to limited expenditures into exposure reduction measures or to a workforce not specifically trained for handling of dangerous chemicals.

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3. Exposure of the consumer When looking at the vast array of chemicals in use today, most of them will have industrial applications and only relatively few will directly be used by consumers. Especially highly reactive chemicals with their inherent potential for health hazards will generally not enter into consumer application fields simply due to their limited stability. But if a chemical gets to the consumer, efficient control measures can hardly be implemented, be it for the exposure per se, the exposed population, appropriate handling or prevention of misuse. 4. Exposure of the general population via the environment Apart from highly reactive substances, all chemicals will to some extent enter the environment depending on their application fields and their processing to other end products. The environmental concentrations are determined by the amount released, by the distribution media, local situations and the efficacy of the different degradation processes. In general, the exposure of the population via the environment will be very low as compared to the workplace and health hazards are not to be expected. Of course highly persistent and highly toxic substances can be an exemption to this rule and should be monitored carefully. These exposure scenarios can be exemplified by a textile dyestuff: manufacturing within the chemical industry makes use of various starting materials and intermediates, which will not end up—apart from minute impurities—in the final product, and downstream exposure to these materials will be negligible. The manufacture of the dyestuff, its further handling, processing and formulation within the chemical industry can easily be controlled. But when this material is used downstream in textiledyeing workshops an efficient exposure control and appropriate handling may not always be guaranteed and higher exposures are conceivable. During the dyeing process, parts of the material enter into the environment, for example into aqueous media. This might lead to an exposure of the general population, albeit at very low concentrations. Consumer exposure can occur via migration of the dyestuff from the textile, sweat being the carrier medium or for small children the saliva, but exposure will most often be so low that a health hazard is not to be expected. It is one of the main concerns of industry within the processes of classification and risk assessment that sufficient consideration is often not given to the different exposure profiles of each chemical. This results in simplified black and white decisions, which are not very helpful for an

H.-P.GELBKE 365

appropriate and cost effective health protection within our industrialized world. Threshold effects Classification The classification system of the EU resides in the designation of appropriate R-phrases. In most cases they are rather meant for hazard identification (e.g. irritation, sensitization, carcinogenicity) and not so much for risk characterization. Sometimes, risk aspects are also involved when specific dose levels are decisive for a specific R-phrase (e.g. acute toxicity). Thereby not only the effect per se is taken into account but also the strength of the effect. This latter principle also applies to the R 48-phrase (‘danger of serious damage to health by prolonged exposure'), if severe (irreversible) toxic effects are observed at dose levels of ≤50 mg kg−1 body weight per day in a 90-day test after oral administration. For other routes and durations of exposure similar provisions exist. The strategy to use dose limits is certainly appropriate for threshold effects. Although for reproductive and developmental effects thresholds are also accepted in most cases, no such dose limits are given apart from 1000 mg kg−1 body weight per day. This dose limit is not equivalent to that for the R 48-phrase, because it only stems from the limit dose levels of the test guideline and is higher by more than one magnitude. Thus, for reproductive/ developmental toxicity classification, exposure and risk considerations are not influential in contrast to the R 48. Such a simplified ‘yes or no’ approach for a threshold effect is not justifiable neither from a scientific point of view nor for an appropriate health protection. This is even more so, if the proposal for the ‘Restrictions on Marketing and Use Directive’ (13th Amendment to Directive 76/769/EEC) is implemented calling for a general prohibition of category 1 and 2 reproductive/ developmental toxicants in consumer products in concentrations of ≥0.5% —if no specific concentration limit has been accepted according to the preparation guideline (Commission Directive 93/18/EEC; Council Directive 88/379/EEC). The inconsistency of the approaches for ‘classical’ organ toxicity and developmental/reproductive toxicity can easily be demonstrated by the following theoretical example: For neurotoxicity a LOEL of 40 mg kg−1 day−1 in a 90-day oral test will result in a R 48-phrase, but a LOEL of 80 mg kg−1 day−1 would not lead to classification. On the other hand, slight foetal weight

366 CONTROVERSIAL ISSUES IN SAFETY ASSESSMENT

reduction or impairment of fertility observed at 800 mg kg−1 day−1 in anappropriate developmental or reproductive test without parental toxicity would lead to classification into the respective category 3 or possibly category 2, even if a NOEL was found at 400 mg kg−1 day−1. This certainly is not appropriate when considering the reversibility and severity of the effects and the differences in the NOELs and LOELs of one order of magnitude. Risk assessment The general strategy in risk assessment of chemicals with threshold effects is to use ‘assessment’ factors (‘uncertainty’ or ‘safety’ factors) (AF) for setting appropriate exposure limits. This concept was originally introduced by the WHO to establish ADI values (acceptable daily intake) for pesticide residues in food. Here generally a ‘safety’ factor of 100 is applied to the NOEL of a chronic experiment; higher or lower factors might be used for specific effects or experimental conditions. This basic approach can be generalized from consumer exposure to pesticide residues in food to other chemicals and exposure scenarios. The main problem then will be the selection of an appropriate AF. First of all, there are some general considerations to be taken into account: – Should AFs for the workforce and the general population differ because of the age characteristics and the general health status of workers? – Should different AFs be selected for chemicals and pesticides, taking into account that pesticides are specifically tailored for biological activity? – What are suitable AFs for route to route extrapolations if the experimental exposure does not correspond to that of humans? Thereby metabolic firstpass effects in the liver and different efficacies of the adsorption barriers of skin, lung and the intestines have to be taken into account. – What is the appropriate dose parameter, mg kg−1 body weight, mg m−2 surface area or concentration? – What are suitable AFs for developmental effects which may occur in principle after a single exposure and may lead to irreversible lifetime impairment? – Are specific AFs necessary for toxic effects on the reproductive organs as compared to toxic effects on other organ systems? Apart from these general considerations there are also specific criteria decisive for the selection of AFs depending on each single chemical, its total data base and the experimental details. Very often the final AF is obtained by additional default factors which are to account for

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experimental insufficiencies or to bridge data gaps. Just to give some examples for such specific considerations: – – – – – – – –

reversible versus irreversible effects, duration of the study, NOAEL versus NOEL, LOAEL versus NOAEL, local versus systemic effects, species-specific effects, species differences in anatomy or physiology, similar versus different results observed in experiments with various species, – biokinetics and metabolism (e.g. metabolic pathways are species specific or occur only at high doses), – structure-activity considerations.

This listing certainly not being complete clearly demonstrates that appropriate AFs cannot be arrived at by a simple cook-book procedure, but a flexible case-by-case approach is required for each individual chemical and data set. This is extremely important in order to avoid overconservative AFs; and in the long run over-conservative risk assessments are just as prohibitive for an appropriate health protection in an industrialized world as an underestimation of risk may lead to a more immediate danger to health. Over-conservative risk evaluations will result in a wrong allocation of resources, an unjustified prohibition of valuable chemicals, a wrong or unnecessary selection of alternative materials. etc. What might be an indication for an over-conservative AF? In principle, AFs for threshold effects should then be questioned to be over-conservative if they lead to acceptable human exposures which would also be appropriate for non-threshold effects (e.g. carcinogenicity) This can be exemplified by the following consideration: For a carcinogenicity experiment a ‘LOAEL’ in classical terms would be equivalent roughly to a dose just leading to a statistically increased tumour incidence of about 5 per cent. A ‘virtual NOAEL’ in the same classical sense without a statistically significant increase could then be at a dose with an actual tumour incidence of 1 per cent, which will not show up as a substance related effect under usual experimental conditions. At such a dose level the extra tumour risk would be 1/100. Applying an AF of 1000 to this ‘virtual NOAEL' would result in an exposure level with a risk of 1/105, and an AF of 10000 in one with a risk of 1/106 using a simple linear extrapolation without further default considerations. Exposure levels with a risk of 1/105 or 1/106 are under discussion as ‘virtually safe doses’ for the workforce or the

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general population. Thus, AFs of >1000 should always be questioned as possibly over-conservative for threshold effects, since exposure levels thereby obtained could also be acceptable for non-threshold effects like carcinogenicity. The assumptions underlying such high AFs should be re-examined critically. In the light of these considerations AFs for a developmental toxicity of 1000– 5000, as they are under discussion by some groups today, should also be questioned as possibly being over-conservative. It should not only be taken into account that developmental effects most often will have thresholds but also that their severities span a wide range from slight foetal weight impairment up to disabling malformations. In addition an unreflected selection of default factors in ‘classical’ organ toxicity can easily lead to over-conservative AFs: starting with a factor of 10 each for inter- and intra-species variability yields the AF of 100 used for ADI-calculations. In addition the following default factors are sometimes proposed: – Extrapolation from subacute/subchronic exposure to chronic exposure: a default factor of 10. – Extrapolation to the NOEL, if only a LOEL was obtained: a default factor of 2–5. – Taking into account an inappropriate experimental design: a default factor of 2–5–10. Thereby, for a multiple dose study with a marginal effect at the lowest dose level and an experimental design not fully in accordance with today’s standards, these default factors would result in a final assessment factor of 4000– 50000. It is highly questionable whether such an AF is really appropriate for threshold effects in comparison to the example given above for carcinogenicity. Non-threshold effects Other principles and approaches for classification and risk assessment have to be applied for non-threshold effects since safe exposure levels cannot be defined at which an adverse health effect will definitely not occur. Thus, for these compounds carcinogenic and mutagenic effects cannot be excluded even at very low dose levels albeit with extremely low probability.

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Classification The present classification systems of scientific organizations (e.g. IARC, German MAK-commission) or regulatory agencies (e.g. EPA, EU commission) reside in quite a simplistic ‘strength of evidence’ approach: how valid are the experimental or epidemiological data? In future we should strive for a ‘weight of evidence’ approach which appears much more appropriate for a realistic human health protection, since it takes into consideration both risks and benefits of man-made and natural chemicals. Such a classification system basically asks the question: what do the experimental data really mean to humans at specific exposure levels? Thereby both qualitative and quantitative aspects are considered. Qualitatively whether and to what extent the mechanisms leading to an adverse effect in animals will also act in humans, and quantitatively to put the experimental dose-response relationship into context with human exposure. Of course, worst case exposure scenarios have to be taken into account. If under these considerations the experimental carcinogenic effect would not be relevant for humans a classification would not be appropriate. This latter quantitative aspect has led to discussions in several groups, as to whether a separate category for ‘weak carcinogens’ should be established, since more and more compounds turn out to be experimental carcinogens with a very weak or questionable effect. Such a category could be used for example for compounds which: – would only have insignificant effects even under worst case exposure scenarios, – did not give a carcinogenic response in appropriate animal experiments but are metabolized to carcinogenic intermediates or exert genotoxic effects in vivo, – show metabolic toxification to genotoxic metabolites only at high doses where the ‘normal’ metabolic detoxification pathway is overwhelmed. It is doubtful whether such a new category would really mean a step forward and be helpful. First of all why should compounds be classified if the carcinogenic effect is not to be expected in humans even under worst case exposure scenarios? And secondly how can the message of ‘weak carcinogenicity’ be brought over to the public without raising an emotional over-reaction to these compounds. Apart from these considerations on the general approach (‘strength’ or ‘weight of evidence’) there is one specific major problem: presently, only criteria for classification are well defined, but those for non-classification are either not or only very vaguely described. It is also an important challenge to set up practicable and clear-cut non-classification criteria

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which can be used by industry to tailor experiments in order to refute the classification of a questionable animal carcinogen. In the long run there will be no benefit if about 50 per cent of all chemicals are classified as carcinogens, neither for the public nor for industry nor for an adequate protection of human health. Risk assessment Most scientific committees and regulatory agencies refrain from scientifically based risk assessments for carcinogens but rather propagate an exposure as low as possible. And if a risk assessment is really carried out, it usually just applies a simplistic mathematical extrapolation using the linearized multistage model and highly conservative default assumptions to bridge data gaps. These mathematical procedures arrive at a scientifically unjustified numerical precision of the risk estimate. One of the problems is to explain to the public the real meaning and the uncertainties of such a risk assessment. A possible alternative could be to substitute the mathematical extrapolation by an appropriate assessment factor which of course has to take into account the severity and irreversibility of the carcinogenic effect. The simplistic mathematical modelling might be used only for selection of priority chemicals for further in-depth investigations. On the other hand, a mathematical risk assessment can be an appropriate procedure for chemicals with a broad experimental data base, when the most relevant default assumptions are substituted by real data. This would be of primary importance for: – the selection of the mathematical model: the simplistic linearized multistage model presently in use could be substituted by biologically driven models, like that proposed by Sielken (1989) or the MVK-model (Moolgavkar and Knudson, 1981; Moolgavkar et al., 1988). – dose scaling from animals to humans: presently the experimental dose in mg kg−1 body weight is often extrapolated to humans by transforming the dose to mg m−2 body surface. This is used both for compounds which are metabolically toxified and detoxified, the scientific basis for such an undifferentiated procedure is at best highly doubtful. In the future this default assumption could be substituted by physiologically based pharmacokinetic (PBPK) modelling. – estimation of the target dose: presently the external dose to which the animals are exposed is considered to be proportional to the dose reaching the target tissue or the target chemical entities—generally the DNA. Again in future this could be substituted by adequate PBPK modelling.

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For the time being, there are only few compounds with an experimental data base broad enough to substitute these default assumptions in every respect. But there are important industrial chemicals—and their number will increase eventually—for which at least some data are available for a justified substitution of part of the default assumptions. And this should be done as far as possible when a mathematical risk assessment is carried out. Conclusions Problems of classification and risk assessment have been discussed separately for toxicological effects with thresholds (‘classical’ organ toxicity, reproductive and developmental toxicity) and without thresholds (mutagenicity, carcinogenicity). With regard to the general procedures of today, for industry there are two main points of concern: 1. Not only for risk assessment but also for classification, exposure considerations should be taken into account. In principle, there are four different exposure scenarios: (a) for chemical production or within chemical industry, (b) for industrial application by downstream users, (c) for the consumer and (d) for the general population via the environment. For a given chemical the exposure may vary widely for the different scenarios, and there are many chemicals which are only used within chemical industry, which will never reach the consumer or which will enter into the environment only in minute quantities. Exposure estimates, including worst case scenarios, should be included in the processes of risk assessment and classification in order to avoid over-conservative results, which are not in the interest of adequate health protection. 2. To get away from risk assessment procedures based on default assumptions which will often lead to over-conservative results; a case by case approach making use of all available data is scientifically far more appropriate. These problems have been elaborated for the: – classification of chemicals for reproductive or developmental toxicity within the regulatory framework of the EU, – selection of appropriate ‘assessment factors’ for chemicals with threshold effects, – classification and risk assessment of chemicals with non-threshold effects (especially carcinogens).

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References MOOLGAVKAR, S.H. and KNUDSON, A.G., 1981, Mutation and cancer: a model for human carcinogenesis, Journal of the National Cancer Institute, 66, 1037–52. MOOLGAVKAR, S.H., DEWANJI, A. and VENZON, D.J., 1988, A stochastic two-stage model for cancer risk assessment, I. The hazard function and the probability of tumour. Risk Analysis, 8, 383–92. SIELKEN, R.L.JR, 1989, Useful tools for evaluating and presenting more science in quantitative cancer risk assessments, Toxic Substances Journal, 9, 353–404.

Index

Absorption of organic solvents 3–7 Acceptable daily intake (ADI) 43, 168 Acetylcholinesterase (AChE) 238 N-acetyl-S-(5-ethoxy-l,2,4-thiadiazol-3yl-methyl) -L-cysteine (ET-MA) 26 N-acetyl-β-glucosaminidase (NAG) 117 Acid anhydrides 152–7 Acrylamide 238 Acrylonitrile 23, 313 Adduct determination 81–7 limitations of methods for 84 Adduct formation of reactive compounds 74 Adipate esters 223 Airway hyperreactivity 128, 131 Airway inflammation, animal models 129 Alachlor 208 Alcohol ethoxylates, anaesthetic properties 349 Alkaline phosphatase (ALP) 117 Alkylating agents 76–9 Alkyldeoxyguanosine adducts 186 Allergic contact dermatitis 137 Allometric equation 44, 46 Allometric scaling 44–55 Allometry 44 Antioxidants 316–39 Aromatic amines 74, 77 Ascorbate 241 Assessment factors (AFs) 366–71, 370 Astrocytes 240 ATP 78, 79, 159 Axonal proteins 238 Azo colorants 301 carcinogenic 301–6

Benzene 39 Benzidene 301 Benzo(a)pyrene (BP) 159, 186 1,4-benzothiazines 64 Benzotriazole-based light stabilisers 322–29 blood kinetics 323–29 blood metabolites 323–29 effects on rat dam and foetal liver 329–4 in vitro hydrolysis 323 liver enzyme induction 327–31 safety assessment 331, 334–7 BGA, collaborative study 199 Bioactivation mechanisms 11, 35–41 Biocides 208 Bioinactivation mechanisms 11 Biological effect monitoring (BEM) 19 Biological effects, determination of 189– 6 Biological monitoring (BM) 19–1 definition 2 organic solvents 2–3 Biomarkers glutathione conjugation products as 20–2 of neurotoxicity 238–4 Biotransformation 12 Bisphenol A diglycidylether (BPADGE) 83 Body metabolic potential 48 Body surface area (BSA) 46–8 BP-DNA adducts 187 Bromobenzene 63 Bromo-diglutathionyl hydroquinones 64

374 INDEX

2-bromo-glutathionyl 64 Bromohydroquinone 63–5 o-bromophenol 63 Bronchial challenge tests 149 Bronchial hyperreactivity 151 Bronchoalveolar lavage 117, 123 1,3-butadiene PBPK/PBTK model 16 physiologically based toxicokinetic modeling 31–3 2-butoxyacetic acid (BAA) 172–80 2-butoxyethanol, PBPK/PBTK model 166–4, 172–80 Calcium 245–9, 250, 267 Cancer risk ‘absolute’ 191 assessment 188–6 definitions 356 management of 181 Carbamazepine 322 Carbendazim 285–8 Carbon disulphide 238 Carcinogen(esis) 179–6, 356 azo colorants 301–6 DNA-reactive 356 dose-response relationship in 356 epidemiological approaches to detect and identify 183–9 genotoxic 181–8 identification of 183–93 peroxisome proliferation 224–30 role of CYPs in activation and detoxication of 206 spontaneous process 356, 357–3 surfactants 350 Carcinogenic potency 181 Carcinogenic Substances Regulation 301 Catabolic metabolism 260 Catalase 241 Cell-mediated immune responses in chemical respiratory allergy 142–6 Cellular nucleophiles 73 Cerebral calcium accumulation 240 Chemical pesticides, immunotoxicity of 203

Chemical respiratory allergy, cellmediated immune responses in 142–6 Chlordecone 241 Chlorinated solvents 223 Chloroform 38–1 Chromates 301 Chrysotile asbestos fibres pulmonary toxic effects 116–30 size-separation methods 118–3 Classification systems 368–3 Color Index 301 Colorants (dyes and pigments) 301–10 regulatory aspects (FRG) 304–7 Cosmetics 208 Cultured porcine thyrocytes 261 Cyclophosphamide 286–90 Cytochrome P-450 (CYP) 15, 60, 74, 206–15, 318, 327, 332 Cytokine products of murine Th1 and Th2 cells 141 Cytokines 140 Dangerous compounds 208 1,2-DCV-Cys 29, 31 DCV-G 29 1,2-DCV-G 29, 31 1,2-DCV-Nac 29 2,2-DCV-Nac 29 Dermal uptake of organic solvents 5–7 Detoxication, enzymes involved in 60–4 Developmental neurotoxins 247–2 Diagonal radioactive zones (DRZ) 158 Diarylide pigments 81–3 1,2-dibromoethane 40, 65–8 3,3′-dichlorobenzidine (DCB) 81–3, 301, 303 1,2-dichloroethane 65–8 2,5-dichloro-3-(glutathion-S-y1) hydroquinone 64 Dichloromethane 39 1,3-dichloropropene (DCP) 21–3 S-(l,2-dichlorovinyl)glutathione. See 1, 2-DCV-G di-(2-ethylhexyl)adipate (DEHA) 223, 224, 227, 229–4 di-(2-ethylhexyl)phthalate (DEHP) 223, 224, 227–4, 286, 287

INDEX 375

5α-dihydrotestosterone 212 Diisocyanate asthma 151 di-(isodecyl)phthalate 223 3,3′-dimethoxybenzine 301 3,3′-dimethylbenzidine 301 Dioctyl phthalate 152 Diphenylhydantoin 322 2,6-di-tert-butyl-4-methyl phenol (BHT) 316–20 DNA adducts 74, 77–81, 83, 84, 158, 159–9, 183–91, 189, 224 immunoenrichment of 185–2 DNA binding 60 DNA damage 183, 192, 208, 356 endogenous 356–1 exogenous 357–3 DNA reactions 73 DNA-reactive carcinogens 356 dose-response curve 356 DNA repair 192 DNA replication 180, 192, 357 DNA synthesis 227, 230 Dopamine 246 Dose determination 188 Dose-response relationship and exposure profiles 362–68 DNA-reactive carcinogens 356 in carcinogenesis 356 low-dose range 190 DTH reactions 196 Dyes. See Colorants (dyes and pigments) EC annex VII and VIII toxicity tests 280 ECETOC 202 EDB, conjugation in rats and man 67 Electrophilic agents 60, 71 Electrophilic centres 72 Electrophilic compounds, examples of 72 Electrophilic metabolites 60 Embryotoxicity tests 289–3 Emulsifiers 312 Endocrine dysfunction, xenobioticinduced 254

Endocrine toxicity, classification of 254– 59 Endocrine toxicology, thyroid 254–80 Entire mammalian tests 283 Environmental exposure 363 Environmental monitoring (EM) 19–1 Epoxide hydrolases (EH) 60 Equivalent radiation dose concept 192– 8 Ethanol 241 5-ethoxy-1,2,4-thiadiazole-3-carboxylic acid (ET-CA) 25–7 Ethylene glycol methyl ether (EGME) 286–91 Etridiazol, disposition of 25–7 European Union (EU) 167 Exposure profiles 362–68 Fatty amines 312 FDA Segment I study for medicines 283 Fecundity tests 284 Fertility and embryotoxicity 285 toxicity 281 Fibre aerosols 97 Fibre glass 91, 99–4 airborne levels in workplace 109 industrial hygiene studies 105–10 lung fibre levels in workers 110 Fibre recovery from lung tissue 118 Fischer 344 study (Kimber-White) 198– 4 Flame retardant 312–18 Flavin-containing monooxygenase (FMO) 211–16 Fluoranthene-DNA adducts 187 Foetal abnormalities 291 Food additives 208 Food and Agriculture Organization 208 Fotemustine 46, 47 Free radical formation 240–4, 246 Full scale testing 288–1 Furazolidone 68 Gas chromatography (GC) 74, 76, 81

376 INDEX

Gas chromatography/mass spectrometry (GC/MS) 74–9, 81, 84, 185 Genotoxic carcinogens 168 Genotoxic hazards 181–8 Genotoxicity 184, 281 in vitro assays 184 Germ cell mutagens 168 Glial fibrillary acidic protein (GFAP) 240–7, 247–2 Gliotypic proteins 238–3 Glutathione (GSH) 15, 20, 24, 184, 241 Glutathione conjugation, reversible 67– 9 Glutathione S-transferase (GST) 15, 20, 24, 60–4, 66, 161, 212, 329 Glycophorin A 160 GM-CSF 143 Growth desensitising mechanism (GDM) 263, 265 Gunn rat hepatocytes in vitro, studies on 274 Haemoglobin 76 Haemoglobin adducts 189 Health surveillance (HS) 19 Heavy metals 244–9 Hepatic metabolism, xenobiotics acting on 267–76 ‘Hepatic pharmacokinetic stuff’ 49 Hepatocarcinogenesis, mechanisms of 226–1 Hepatocytes co-cultures phase 1 reactions 210–16 phase 2 reactions 212–17 long-term cultures of 207–14 Herbicides 208 Hexamethylene diisocyanate (HDI) 151 n-hexane 3, 7, 238 2,5-hexanedione 3 HGPRT mutation assay 192 HHPA 152 Himic anhydride (HA) 152 Hormone elimination 269 Hormone synthesis 258

Host resistance (HR) studies 200 HPLC 74–83, 84 Human allergic disease 142 Human clearance prediction 49, 53 Human exposure monitoring 188 Human unbound clearances 51–3 Hydroxymethylethenodeoxyadenosine (HMEdA) 83 1-hydroxypyrene 159 Hypersensitivity reactions, Type I-IV 196 Hypothalamic-pituitary-thyroid-liver (H-P-T-L) axis 256–60 investigative tests on 267 thyroid toxicity via 265–76 toxicological 266 xenobiotic toxic effects on 258 Hypoxanthin guanine phosphoribosyl transferase (HPRT) 160 ICICIS collaborative study 197–3 IFN- 140–6 IgA 199, 200 IgE 153 IgE antibody 138 IgE antibody responses, induction and regulation of 140–5 IgG 153, 200 IgG2a antibody 140 IgM 200 Immune system, evaluation of toxicity to 196–10 Immunoenrichment of DNA adducts 185–2 Immunological methods 77, 78 Immunotoxic side effects, screening of 197 Immunotoxicity of chemical pesticides 203 Immunotoxicity testing, direct food additives 202 Immunotoxicology 196 collaborative studies 197, 198 screening tests 196 Insulin-like growth factor 1 (IGF,) 265 Interferon (IFN- ) 140–6 Interleukin 3 (IL-3) 141,143

INDEX 377

Interleukin 4 (IL-4) 140, 141, 143 Interleukin 5 (IL-5) 141–6 Interleukin 10 (IL-10) 141 Interleukin 12 (IL-12) 142 International Agency for Research on Cancer (IARC) 91, 93–7, 163 International Programme on Chemical Safety (IPCS) 91 Iron 246 Isocyanates 151–6 Isophorone diisocyanate (IPDI) 151 Joint Expert Committee on Food Additives (JECFA) 208 Kainic acid (KA) 242–7 Labelling 296–9 Lactate dehydrogenase (LDH) 117 Late respiratory systemic syndrome (LRRS) 152–7 Lead 244–8 Leaving group 65–8 Life cycle exposure 283 Life span correction 47–49 Light microscopic histopathology 123–6 Light stabilisers 316–39 benzotriazole-based 322–9 Linearized multistage cancer model (LMS) 170 Lipophilic compounds 206 Liver enzyme induction 318–2 benzotriazole-based light stabilisers 327–31 LOAEL 168, 367 Local lymph node assay (LLNA) 196 LOEL 43 Low molecular weight (MW) organic chemicals 187 Lung burden analysis 103–8, 121–5 Lung digestion/biodurability studies 125 Lung dissection 118 Lung fibre burden 98–1 Lung tissue, fibre recovery from 118 MAK-list 304

Maleic anhydride (MA) 152 Malformations 291 Malignancy, critical mutations leading to 190 Manganese 245, 246 Man-made vitreous fibres (MMVFs) animal inhalation studies 95–106 carcinogenic potential 91–115 cell culture studies 94 comparison of human exposures used in rat chronic inhalation studies 108–13 epidemiological studies 91, 93–7 implantation studies 95 potential biological effects 91 previous inhalation studies 106–9 toxicologic studies 94 Maximum life potential (MLP) 47–54 Maximum tolerated dose (MTD) 170 Meehs Formula 47 Mercaptans 21 Mercapturic acids 21–4, 184–90, 187–3 toxicokinetics of 23 urinary excretion 15, 25 Metabolic activation 60–4 Metabolism and toxicity 206–12 Methamphetamine 241 Methylene chloride 65 Methylene diphenyldiisocyanate (MDI) 151 Methylmercury 241, 245 N-methyl-N-nitrosourea (MNU) 263 Michaelis-Menten kinetics 173 Mitogenic stiraulation (ConA.LPS) 199 Model neurotoxins 241–52 Monitoring environmental (EM) 19–1 human exposure 188 in occupational toxicology 19–1 polycyclic aromatic hydrocarbon (PAH) exposure 158–6 see also Biological effect monitoring (BEM); Biological monitoring (BM) Monoclonal antibodies (Mabs) 185–2, 318 MPP+ 241 MPTP 241

378 INDEX

mRNA analysis 211 Mutagenic potency 190–6 Mutagenicity, surfactants 349–2 N7-deoxyguanosine (N7-dG) 186 NADPH-cytochrome P450 reductase 15 Naphthalene diisocyanate (NDI) 151 β-naphthoflavone 266, 267 2-naphthylamine 40–3, 301 Neuropathy target esterase (NTE) 238 Neurotoxicity, biomarkers of 238–4 Neurotoxicity assessment 285 Neurotoxicity testing 237–56 Neurotypic proteins 238–3 Nitroarenes 74, 77 NK activity 199 NK test 200 NOAEL 168, 174, 367 NOEL 43, 366, 367 Nucleophilic centres 73 Occupational asthma 148–9 chemical agents causing 148–5 incidence 148, 151 initial diagnosis 149 Occupational toxicology, monitoring in 19–1 OECD guidelines 421 282–7 OECD guidelines 422 282–91 OECD single generation study 283 17β-oestradiol 212 Organic solvents absorption of 3–7 biological monitoring of 2–3 dermal uptake of 5–7 pulmonary uptake of 3–5 Organophosphate-induced delayed neuropathy (OPIDN) 238 Organophosphate pesticides 237–2 Parkinson’s disease 241, 245 PBPK/PBTK models 1,3-butadiene 16 2-butoxyethanol 166–4, 172–80 development of 31–3 in risk assessment 170–80

PCA-DNA adducts 187 PCNB 267 Pentafluorophenyl isothiocyanate (PFPITC) 76 Pentafluorophenyl thiohydantoine (PFPTH) 77 Perchlorate-discharge test 260 Peroxisome proliferation 223–40 carcinogenicity 224–30 in rodent liver 223–8 mechanisms of 226–1 risk assessment 229–5 rodent 223–9 species differences in response 227–3 Pesticides 203, 202 human exposure to 208 immunotoxicity of 203 organophosphate 237–2 Pharmaceuticals 202 Phencyclidine 48 Phenobarbital 267, 273 Phenobarbitone 322 Phenolic antioxidants 316–25 blood kinetics 317–1 blood metabolites 317–1 effects on serum thyrotropin and thyroid hormones 318–4 liver enzyme induction 318–2 model compound 317 risk assessment 321–5 Phthalate esters 223, 224 Phthalic anhydride 152 Physiologically-based pharmaco (toxico-) kinetic models. See PBPK/ PBTK models Phytopharmaceuticals 208 Pigments. See Colorants (dyes and pigments) Plaque assay (PFCA) 199, 200 Plasticisers 223, 229–4 Polychlorinated biphenyls (PCBs) 247– 2, 267 Polycyclic aromatic hydrocarbons (PAHs) 74–7, 78, 158, 186 biomonitoring exposure 158–6 Polyisocyanates 151–6 Postlabelling 78–79, 81 Post-natal manifestations 284

INDEX 379

Post-radiolabelling technology 184 Prenatal effects 284, 285, 289 Production volume triggers 280, 281 Propylthiouracil (PTU) 260 Protein adducts 73, 74, 79, 81, 184–90 Protein-binding 258 Pulmonary cell proliferation 118, 124–7 Pulmonary hyperreactivity to industrial pollutants 128–9 guinea-pig model 131–6 rat model 134–7 Pulmonary sensitization, mechanisms of 137–51 Pulmonary toxic effects chrysotile asbestos fibres 116–30 para-aramid fibrils 116–30 Pulmonary uptake of organic solvents 3–5 Pyrethroid insecticides 238 Pyromellitic dianhydride (PMDA) 152 Quinone-thioethers 63 Quinones 63–6 Rad-equivalence values 192 Radioallergosorbent tests (RAST) 138 Radioimmunoassays 77 Reactive airways dysfunction syndrome (RADS) 128, 148, 151, 153 Reactive chemicals, metabolism 60–71 Reactive compounds adduct formation of 74 determination in unknown mixtures 83–6 determination of 71–88 interaction with cellular constituents 71–5 Reactive metabolites 71 Refractory ceramic fibres (RCFs) 91, 99–2 industrial hygiene studies 105 Repeated dose toxicity 281 Reproductive toxicity 279–298 detecting effects on males 293 evaluation for 282 interpretation/extrapolation of 296– 9

labelling 296–9 manifestations of 291–4 methods for detecting effects on 282 overlap of 281 Respiratory allergic hypersensitivity 137 Restricted test systems 283 Rifampicin 322 Risk assessment 162–9, 166–84 and exposure profiles 362–68 approaches to 167–6 cancer 188–6 guidelines for 361 in vitro approach 207–13 mathematical procedures 369–3 non-threshold effects 368 peroxisome proliferation 229–5 phenolic antioxidants 321–5 physiologically based pharmacokinetic (PBPK) models in. See PBPK/PBTK models textile chemicals 315 threshold effects 366–71 see also Safety assessment RNA probes 211 Rock (stone) wool 91, 102–5 industrial hygiene studies 108 R-phrases 364 Safety assessment 361–74 benzotriazole-based light stabilisers 331–7 see also Risk assessment Salmonella typhimurium 66 SHIELD system 149 Slag wool 91, 102–5 industrial hygiene studies 108 Solid phase assays 77 Sperm analysis 284 Spermatogenesis 284, 286, 288 Structure-activity databases 282 Structure-activity relationships, surfactants 344 Styrene 38–38 Styrene oxide 38–38 Superoxide dismutase (SOD) 241 Surface markers 200

380 INDEX

Surfactants 338–55 acute toxicity 348–1 biochemical properties 338–5 carcinogenicity 350 chronic toxicity 349 embryotoxicity 350 excretion 347–50 interactions with enzymes 341–5 interactions with membranes 339 interactions with proteins 339–3 intestinal absorption 347–50 local effects 342–48 metabolism 347–50 mucous membrane compatibility 345 mutagenicity 349–2 oral toxicity 348 percutaneous absorptions 346–9 sensitization 345–8 skin compatibility 342–7 structure/activity relationships 344 systemic effects 348 toxicokinetics 346–50 Surveillance of Work-related and Occupational Respiratory Disease Project (SWORD) 148 Synaptophysin 242 Systemic bioavailability of colorants 301 Systemic toxicity assessment 285 Target dose determination 188–4 Temelastine 271 272, 273 2-tert-buty1–4-methoxyphenol (BHA) 316 Testosterone 212 Tetrachloroethane 7 Tetrachloroethene 8, 7 Tetrachlorophthalic anhydride (TCPA) 152, 153 Textile chemicals 308–18 finishing plant 309–13 handling and processing 312–18 irritant properties 311, 312 new developments regarding toxicology 309 oral toxicity 310

process off-gas 315 risk assessment 315 temperature effects 311 toxicological profile 311 toxicology assessment needs 315 Thalidomide 291 T helper (Th) cells 140–5 Thioethers 21–3 Threshold effects classification 364–9 risk assessment 366–71 Thyroglobulin (TBG) 258, 260 Thyroid endocrine toxicology 254–80 tumours of 256 Thyroid binding pre-albumin (TBPA) 258–2 Thyroid follicles 260 Thyroid follicular capability 262 Thyroid follicular cell hyperplasia and neoplasia, pathobiology of 262–7 Thyroid function control of 258 perturbation of 256–60 Thyroid hormones 258, 260, 318–4 Thyroid lesions, pathobiology of 258 Thyroid neoplasia 262, 321 Thyroid stimulating hormone (TSH) 258, 260, 262, 263, 265, 320, 321, 322 Thyroid toxicity 256 via H-P-T-L axis 265–76 Thyroid tumorigenesis 264 Thyrotrophin releasing hormone (TRH) 258, 262 Thyrotropin 318–4 Thyroxine 260, 269 accumulation 271, 273 clearance of 267–76 T lymphocytes 142 Toluene 241 2,4- and 2,6-toluene diisocyanate (TDI) 151 Toxicity, and metabolism 206–12 Toxicity tests, EC annex VII and VIII 280 Toxicodynamic interactions 20 Toxicokinetic interactions 20

INDEX 381

Toxicokinetic parameters 16 Toxicokinetics, principles of 15 Transition metals 246 Trichloroacetic acid (TCA) 28, 29 Trichloroethane 7, 8 Trichloroethylene (TRI) 26–31 effects related to 26–31 hepatocarcinogenicity induced by 28 oxidative metabolism 28, 29 2,5,6-trichloro-3-glutathion-S-yl) hydroquinone 64 Triethyl lead 241 tri-(2-ethylhexyl)trimellitate) 223 Trimellitic anhydride (TMA) 152, 153 Trimethyltin (TMT) 241–6 Tumour incidence 189–5 Two generation studies 291–5 alternatives to 295–8 UDP-glucuronosyltransferase 321, 327, 329 UDP-glucuronyltransferase (UDP-GT) 213, 260, 267, 271, 274 University of Pittsburgh study 91, 92 Urinary excretion mercapturic acids 15, 25 xenobiotics 16–18 US Environmental Protection Agency 91 US-NTP study 200 Veterinary medicinal chemicals 202 Vinyl chloride 36–9 Western blotting 211 World Health Organization (WHO) 91 Xenobiotic-induced endocrine dysfunction 254 Xenobiotics 11 assessment of long-term toxicity 207 biological effects 14 disposition of 12–15 overall exposure to 20 urinary excretion of 16–18