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 experimental